a strategy for xenobiotic removal using - infoscience
TRANSCRIPT
POUR L'OBTENTION DU GRADE DE DOCTEUR ÈS SCIENCES
PAR
Ingénieur diplômée de l'ENS de Biologie Appliquée à la Nutrition et à l'Alimentation, Université de Dijon, Franceet de nationalité française
acceptée sur proposition du jury:
Lausanne, EPFL2006
Prof. A. Mermoud, président du juryDr C. Pulgarin, directeur de thèse
Dr L. De Alencastro, rapporteurDr S. Malato Rodriguez, rapporteur
Prof. C. Petrier, rapporteur
a strategy for xenobiotic removal using photocatalytic treatment,
microbial degradation or integrated photocatalytic-biological process
Miléna LAPERTOT
THÈSE NO 3548 (2006)
ÉCOLE POLYTECHNIQUE FÉDÉRALE DE LAUSANNE
PRÉSENTÉE LE 23 jUIN 2006
à LA FACULTÉ ENVIRONNEMENT NATUREL, ARCHITECTURAL ET CONSTRUIT
Laboratoire de biotechnologie environnementale
SECTION DES SCIENCES ET INGÉNIERIE DE L'ENVIRONNEMENT
Miléna Lapertot PhD. Thesis
I
Summary
According to the limited natural resources and due to the risks of anthropogenic pollution, it
appears necessary to react efficiently in order to remove existing contaminations and avoid
the creation of new ones. Therefore, the purpose of this thesis is to propose a sustainable
strategy for treating problematic pollutants with the most adequate process.
First, an overview of the different treatment processes has been given. In particular,
biological, photocatalytic and integrated biological-photocatalytic treatments have been
described as efficient methods to remove xenobiotics. Biological treatment of contaminated
environments and industrial effluents has been presented as the most attractive method on a
cost-benefit extent. But for biorecalcitrant substances, the Advanced Oxidation Processes
(AOP) have remained the most efficient methods although expensive. Therefore, the
coupling of biological-photocatalytic has emerged as the best compromise for removal of
recalcitrant xenobiotics.
The first step of the treatment strategy has consisted of evaluating the biotraitability of 19
chemicals: pesticides, pharmaceuticals and volatile organic compounds (VOCs) have been
assessed according to three biodegradability methods (Zahn-Wellens, Manometric
Respirometry, Closed-Bottle tests), whose official protocols had to be adapted to the
compound specificities (volatility, hydrophobicity). Structure-Activity Relationship (SAR)
estimation models have also been used to compare and validate the experimental results.
For the compounds which have been assessed as biotraitable (VOCs), further biodegradation
studies have been led in order to improve the treatment (batch with pulses of substrate):
automated monitoring of the biodegradation process, increased degradation yields and
biomass productivity have been successfully completed for Toluene, Ethylbenzene, Xylenes,
Chlorobenzenze and Dichlorobenzene removal.
Conversely, the compounds assessed as biorecalcitrant have been oriented towards the TiO2
and photo-Fenton photocatalytic treatments. Thus, the degradation of four p-halophenols by
TiO2 photocatalysis has been investigated and has demonstrated the halogen effect on
removal rates, aromatic intermediates and toxicity variations. Finally, the treatment of five
Miléna Lapertot PhD. Thesis
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bioreacalcitrant pesticides (Alachlor, Atrazine, Chlorfenvinphos, Diuron, Pentachlorophenol)
has been studied in order to develop a coupled photocatalytic-biological system: first,
enhanced biotreatability has been evaluated using the Zahn-Wellens procedure and then
fixed-bed bioreactors have been operated and permitted to find out the best moment for
coupling.
Key words: biodegradability, xenobiotics, photocatalysis, pollution removal, wastewater,
fed-batch, pesticides, VOCs.
Résumé
En raison de la limitation des ressources naturelles et des risques de pollutions
anthropogéniques de l’environnement, il apparaît nécessaire d’agir efficacement en vu
d’évincer les contaminations existantes et d’éviter la dissémination de nouvelles sources
polluantes. Le but de cette thèse est alors de proposer une stratégie permettant de déterminer
le processus le plus adéquat pour traiter des polluants problématiques.
Un survol des différents procédés de traitement a d’abord été présenté. L’efficacité des
méthodes biologiques, photocatalytiques et de couplage biologique-photocatalytique a alors
été établie pour l’élimination des composés xénobiotiques. Le traitement biologique des sites
contaminés et des effluents industriels s’est imposé comme la méthode la plus attractive en
terme de coûts-bénéfices. Néanmoins, les procédés d’oxydation avancée (POA) s’avèrent
encore être les plus efficaces pour traiter les substances biorécalcitrantes, bien qu’ils
occasionnent des coûts élevés. De ce fait, le couplage biologique-photocatalytique semble
être le meilleur compromis pour l’élimination des composés xénobiotiques récalcitrants.
La première étape de la stratégie de traitement établie dans le cadre de cette thèse a consisté a
évaluer à la biotraitabilité de 19 produits chimiques : pesticides, composés pharmaceutiques
et composés organiques volatils (COVs) ont été évalués sur la base de trois méthodes (Zahn-
Wellens, Respirometrie Manometrique, Tests des bouteilles fermées), dont les protocoles
officiels ont dû être adaptés aux spécificités des composés étudiés (volatilité,
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hydrophobicité). Des modèles de relation « Structure-Activité » ont également été employés,
pour comparaison et validation des résultats expérimentaux.
Pour les composés évalués biotraitables (COVs), des études de biodégradation ont été
menées dans l’optique d’améliorer le procédé de traitement (système batch avec pulses de
substrat) : le monitoring automatisé du procédé de dégradation, l’augmentation des
rendements de dégradation et de la productivité de biomasse ont été réalisés avec succès pour
l’élimination du toluène, de l’ethylbenzène, des xylenes, du chlorobenzène et du
dichlorobenzène.
Au contraire, les composés évalués biorécalcitrants ont été orientés vers un traitement
photocatalytique par TiO2 ou photo-Fenton. La photodégradation de quatre p-halophénols par
TiO2 a été investiguée et a permis de démontrer l’effet de l’halogène sur les taux
d’élimination du polluant, la nature des intermédiaires aromatiques produits et les variations
de toxicité de la solution traitée. Enfin, un système de couplage biologique-photocatalytique
a été développé et étudié pour le traitement de cinq pesticides biorécalcitrants (Alachlor,
Atrazine, Chlorfenvinphos, Diuron, Pentachlorophénol) : tout d’abord, une biotraitabilité
améliorée des pesticides a été réalisée et évaluée selon le protocole du test de Zahn-Wellens,
puis leur traitement en bioréacteurs sur lits fluidisés a été accompli et a permis de déterminer
le moment le plus approprié au couplage.
Mots-clés: biodegradabilité, xénobiotiques, photocatalyse, dépollution, eaux résiduaires, fed-
batch, pesticides, COV.
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Table of contents
Summary / Résumé .............................................................................................................I
Table of contents...............................................................................................................V
List of Figures..................................................................................................................XI
List of Tables ................................................................................................................. XV
List of Schemes............................................................................................................XVII
CHAPTER 1..........................................................................................1
INTRODUCTION AND BACKGROUND OF THE THESIS
1.1. Context..........................................................................................................................1
1.2. Conventional treatments for pollution removal .............................................................3
1.2.1. Biological treatments .............................................................................................3 1.2.2. Conventional wastewater treatments ......................................................................9 1.2.3. Advanced Oxidation Processes (AOP).................................................................11
1.2.3.1. Heterogeneous photocatalysis .......................................................................13 1.2.3.2. Homogeneous photodegradation ...................................................................15 1.2.3.3. Fenton and photo-Fenton assisted reactions ..................................................16
1.3. Aim of the thesis .........................................................................................................18
1.4. The target xenobiotics compounds..............................................................................20
1.4.1. Volatile Organic Compounds...............................................................................20 1.4.2. Phenols.................................................................................................................22 1.4.3. Pesticides..............................................................................................................24
1.5. References...................................................................................................................26
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CHAPTER 2. ......................................................................................37
BIODEGRADABILITY ASSESSMENT OF SEVERAL PRIORITY HAZARDOUS SUBSTANCES: CHOICE, APPLICATION AND RELEVANCE REGARDING TOXICITY AND BACTERIAL ACTIVITY
2.1. Abstract.......................................................................................................................37
2.2. Introduction.................................................................................................................37
2.3. Material and methods..................................................................................................40
2.3.1. Chemicals.............................................................................................................40 2.3.2. Biodegradability methods ....................................................................................42
2.3.2.1. Zahn-Wellens test..........................................................................................42 2.3.2.2. Closed bottle and Manometric respirometry tests .........................................42
2.3.3. Initial concentrations ............................................................................................43 2.3.4. Analytical methods...............................................................................................43
2.4. Results and discussion.................................................................................................45
2.4.1. Selecting the appropriate test of biodegradability ................................................45 2.4.2. Zahn-Wellens test.................................................................................................47 2.4.3. “Ready-biodegradability” tests.............................................................................49 2.4.4. Predictive estimation models................................................................................53
2.4.4.1. Qualitative Substructure Model.....................................................................53 2.4.4.2. Biodegradability Probability Program...........................................................54
2.5. Conclusion ..................................................................................................................57
2.6. References...................................................................................................................57
CHAPTER 3........................................................................................61
BIODEGRADATION OF VOLATILE ORGANIC COMPOUNDS
PART A.............................................................................................................................61
Enhancing production of adapted bacteria to degrade chlorinated aromatics
3.A.1. Abstract ...................................................................................................................61
3.A.2. Introduction.............................................................................................................62
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3.A.3. Materials and methods ............................................................................................63 3.A.3.1. Inoculum Preparation........................................................................................63 3.A.3.2. Bioreactor and Automation Program................................................................64 3.A.3.3. Analytical methods ...........................................................................................65
3.A.4. Results and discussion.............................................................................................67
3.A.4.1. Substrate Feeding Pattern .................................................................................67 3.A.4.2. Parameters of Control .......................................................................................69
3.A.4.2.1. Metabolic intermediates ............................................................................69 3.A.4.2.2. Chloride concentration ..............................................................................72 3.A.4.2.3. Nitrogen contents ......................................................................................73 3.A.4.2.4. Cellular Integrity .......................................................................................73
3.A.4.3. CB and DCB Conversion..................................................................................74 3.A.4.3.1. SPB Technique..........................................................................................74 3.A.4.3.2. Continuous Feeding...................................................................................76
3.A.4.4. Biomass Yield and Productivity .......................................................................77
3.A.5. Conclusion ..............................................................................................................79
3.A.6. Nomenclature ..........................................................................................................80
3.A.7. References ...............................................................................................................80 CHAPTER 3........................................................................................83
BIODEGRADATION OF VOLATILE ORGANIC COMPOUNDS
PART B.............................................................................................................................83
Mass production of bacterial communities adapted to the degradation of volatile organic compounds (TEX)
3.B.1. Abstract ...................................................................................................................83
3.B.2. Introduction .............................................................................................................84
3.B.3. Materials and methods.............................................................................................85
3.B.3.1. Inoculum Preparation........................................................................................85 3.B.3.2. Bioreactor and Automation Program................................................................86 3.B.3.3. Analytical methods ...........................................................................................87
3.B.4. Results and discussion.............................................................................................89
3.B.4.1. Bioengineering..................................................................................................89 3.B.4.1.1. SPB technique for biomass cultivation ......................................................89 3.B.4.1.2. Automation parameters..............................................................................91
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3.B.4.2. Microbiology ....................................................................................................93 3.B.4.2.1. Microbial growth .......................................................................................94 3.B.4.2.2. Impact of C and N concentration on cell viability .....................................95 3.B.4.2.3. Substrate conversion..................................................................................96 3.B.4.2.4. Substrate uptake rate..................................................................................99
3.B.5. Conclusion.............................................................................................................100
3.B.6. Nomenclature ........................................................................................................101
3.B.7. References .............................................................................................................101
CHAPTER 4......................................................................................105
PHOTOCATALYTIC DEGRADATION OF P-HALOPHENOLS IN TiO2 AQUEOUS SUSPENSIONS: HALOGEN EFFECT ON REMOVAL RATE, AROMATIC INTERMEDIATES AND TOXICITY VARIATIONS
4.1. Abstract.....................................................................................................................105
4.2. Introduction...............................................................................................................106
4.3. Materials and methods ..............................................................................................107
4.3.1. Materials ............................................................................................................107 4.3.2. Adsorption isotherms in the dark .......................................................................107 4.3.3. Photocatalytic degradation experiments .............................................................107 4.3.4. Analytical methods.............................................................................................108
4.4. Results and discussion...............................................................................................110
4.4.1. Adsorption of phenol and p-halophenols on TiO2 ..............................................110 4.4.2. Photocatalytic removal rates of phenol and p-halophenols ................................113
4.4.2.1. Results.........................................................................................................113 4.4.2.2. Correlation between the photocatalytic degradability and the adsorbed amounts of p-halophenols ........................................................................................114 4.4.2.3. Correlation between the photocatalytic degradability and the Hammett constant of p-halophenols.........................................................................................114
4.4.3. Aromatic intermediate products of degradation .................................................115 4.4.4. Toxicity ..............................................................................................................120 4.4.5. Carbon balance...................................................................................................123
4.5. Conclusion ................................................................................................................123
4.6. References.................................................................................................................124
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CHAPTER 5......................................................................................127
INTEGRATED PHOTOCATALYTIC-BIOLOGICAL PROCESS FOR TREATMENT OF PESTICIDES
PART A...........................................................................................................................127 Evaluating Microtox© as a tool for biodegradability assessment of partially treated solutions of pesticides using Fe3+ and TiO2 solar photo-assisted processes
5.A.1. Abstract .................................................................................................................127
5.A.2. Introduction...........................................................................................................128
5.A.3. Experimental .........................................................................................................130 5.A.3.1. Chemicals .......................................................................................................130 5.A.3.2. Analytical determinations ...............................................................................131 5.A.3.3. Toxicity Measurements...................................................................................131 5.A.3.4. Biodegradability assessment ...........................................................................132 5.A.3.5. Photoreactor set-up .........................................................................................132 5.A.3.6. Data analyses ..................................................................................................134
5.A.4. Results...................................................................................................................135
5.A.4.1. Photodegradation ............................................................................................135 5.A.4.2. Toxicity Measurements...................................................................................139 5.A.4.2. Biodegradability testing ..................................................................................141
5.A.5. Discussion .............................................................................................................143
5.A.5.1. Toxicity assessment ........................................................................................143 5.A.5.2. Biodegradability Assessment..........................................................................144
5.A.6. Conclusions ...........................................................................................................146
5.B.7. References .............................................................................................................147
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CHAPTER 5......................................................................................153
INTEGRATED PHOTOCATALYTIC-BIOLOGICAL PROCESS TO TREAT PESTICIDES
PART B...........................................................................................................................153
Photo-Fenton and biological integrated process for degradation of a mixture of pesticides
5.B.1. Abstract .................................................................................................................153
5.B.2. Introduction ...........................................................................................................154
5.B.3. Materials and methods...........................................................................................155
5.B.3.1. Chemicals .......................................................................................................155 5.B.3.2. Analytical determinations ...............................................................................156 5.B.3.3. Toxicity measurements ...................................................................................156 5.B.3.4. Biodegradability assessment ...........................................................................157 5.B.3.5. Experimental set-up ........................................................................................157
5.B.3.5.1. Photo-Fenton experiments .......................................................................157 5.B.3.5.2. Biotreatment experiments ........................................................................158
5.B.3.6. Data analyses ..................................................................................................159
5.B.4. Results and discussion...........................................................................................160 5.B.4.1. Biocompatibility of the mixed pesticide solution............................................160 5.B.4.2. Chemical and biological characteristics of the photo-treated pesticide solution.....................................................................................................................................161 5.B.4.3. Photochemical-biological coupled flow treatment ..........................................166
5.B.5. Conclusion.............................................................................................................169
5.B.6. References .............................................................................................................169
CHAPTER 6......................................................................................173
CONCLUSION AND PERSPECTIVES OF THE THESIS
Acknowledgments ........................................................................................................179
Curriculum Vitae .................................................................................181
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List of Figures
Figure1.1. Main principle of aerobic degradation of hydrocarbons................................... 5
Figure 1.2. Suitability of water treatment technologies according to the Chemical Oxygen
Demand, COD. ........................................................................................................ 12
Figure 1.3. Main advanced oxidation processes (AOP). ................................................. 13
Figure 1.4. Schematic of an irradiated TiO2 semi-conductor particle with some possible
photo-chemical and photo-physical processes. ........................................................ 15
Figure 1.5. Overview of the strategy proposed to treat xenobiotics ................................ 19
Figure 2.1. Zahn-Wellens biodegradability test. Evolution of the degradation percent as a
function of time for R113675, T824, FEMAM, diuron, isoproturon and diethylene
glycol DEG (control). .............................................................................................. 48
Figure 2.2a. Evolution of the VOC degradation during 28 days for the Ready-
Biodegradability testing (BOD28)............................................................................. 50
Figure 2.2b. Evolution of the degradation percent during 28 days for the Ready-
Biodegradability testing of Alachlor, Chlorfenvinphos (Manometric Respirometry),
Atrazine, Lindane and Pentachlorophenol (BOD28)................................................. 50
Figure 3.A.1. Schematic presentation of the laboratory bioreactor ................................. 65
Figure 3.A.2. Disappearance of CB and DCB in the liquid phase, consumption of Dissolved
Oxygen, and evolution of absorbance at 255 nm during one pulse. ......................... 68
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Figure 3.A.3. Calculation and evolution of the degradation rates qs during the cultivation
process ..................................................................................................................... 76
Figure 3.B.1. Schematic presentation of the laboratory bioreactor ................................. 87
Figure 3.B.2. Evolution of DO during automatic runs of biomass cultivation ................ 93
Figure 3.B.3. Evolution of the biomass concentration during the cultivation process..... 94
Figure 3.B.4. Disappearance ratio of TEX, oxygen uptake and A255 evolution during a
representative pulse, after 4 days of cultivation. ...................................................... 97
Figure 3.B.5. Disappearance of TEX checked along 30 days of cultivation ................... 98
Figure 3.B.6. Evolution of the degradation rates qs during the cultivation process ....... 100
Figure 4.1. Room temperature isotherms of adsorption on TiO2 for phenol and p-
halophenols.. .......................................................................................................... 111
Figure 4.2. Kinetics of the photocatalytic disappearance of phenol and p-halophenols. The
inset shows linear transforms.. ............................................................................... 113
Figure 4.3. Percentages of 4-chloro-1,3-dihydroxybenzene and 4-chloro-1,2-
dihydroxybenzene formed with respect to the initial pollutant concentration versus the
percentage of pollutant degraded. .......................................................................... 116
Figure 4.4. Percentages of BQ, HQ and BQ + HQ (with respect to the initial pollutant
concentration) versus the removed percentage of phenol and p-halophenols......... 117
Figure 4.5. Evolution of the acute toxicity during the photocatalytic degradation of phenol
and p-halophenols. ................................................................................................. 120
Figure 4.6. Percentages of the relative carbon forms in the TiO2 suspensions during the
photocatalytic degradation of p-halophenols.......................................................... 122
Figure 5.A.1. Flow diagram of photoreactor ................................................................. 133
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Figure 5.A.2. Evolution of mineralization, compound removal and toxicity during the TiO2
photocatalytic process of single pesticide solutions tested .................................... 137
Figure 5.A.3. Evolution of mineralization, compound removal and toxicity during the photo-
Fenton process of single pesticide solutions tested ................................................ 138
Figure 5.B.1. Experimental setup used for investigation of the photocatalytic-biological
coupling treatment of the mix of pesticides. .......................................................... 159
Figure 5.B.2. DOC evolution of the solution of mixed pesticides as a function of the
irradiation time (t30W) and the consumed amounts of H2O2 during photo-Fenton. . 162
Figure 5.B.3. Evolution of toxicity and DOC of the solution of the mixed pesticides as a
function of the irradiation time (t30W) during photo-Fenton. .................................. 163
Figure 5.B.4. Relative contribution of photocatalysis and biotreatment for the degradation of
the solution of the mixed pesticides. ...................................................................... 168
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List of Tables
Table 1.1. Beneficial and deleterious effects of biodegradation ........................................ 6
Table 1.2 Oxidation potential of several oxidants in water ............................................ 11
Table 2.1. Physico-chemical properties of the target compounds ................................... 41
Table 2.2. Test methods and results for the biodegradability assessment of the target
compounds.. ............................................................................................................. 46
Table 2.3. Comparison of experimental results and estimated data for the biodegradability
assessment of the 19 xenobiotics ............................................................................. 55
Table 3.A.1. Operational data for the bioreactor ............................................................. 64
Table 3.A.2. Production of intermediate metabolites as the bacterial response to continuous
or batch cultivation processes. ................................................................................. 70
Table 3.A.3. Biomass yields and productivities achieved by cultivation processes operated
within aerated SPB, oxygenated SPB or continuous feeding techniques. ................ 78
Table 3.B.1. Operational data for the bioreactor ............................................................. 86
Table 3.B.2. List of the National Instruments modules ................................................... 86
Table 3.B.3. Influence of C and N on the cellular metabolism of the mixed population. 96
Table 4.1. Measured extents of adsorption, molecular characteristics and measured values of
k for phenol and p-halophenols. ............................................................................. 112
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Table 5.A.1. Experimental and reference data for acute Microtox© EC50 testing for the target
pesticides................................................................................................................ 140
Table 5.A.2. Biodegradability assessment performed with the partially phototreated solutions
of single pesticide. ................................................................................................. 142
Table 5.B.1. Physico-chemical and biological properties of the pesticides. .................. 161
Table 5.B.2. Comparison of the illumination time (t30W) necessary to remove 100% of the
parent molecule of pesticide and 80% of DOC during the photo-Fenton treatment of
both single and mixed pesticides............................................................................ 164
Table 5.B.3. Photocatalytic characteristics and biodegradability assessment of the pre-treated
solutions tested with the photo-Fenton / Bioreactor coupled system...................... 166
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List of Schemes
Scheme 1.1. Structures of the VOCs studied in Chapters 2 and 3. .................................. 21
Scheme 1.2. Structures of the phenolic substances studied in Chapters 2 and 4.............. 23
Scheme 1.3. Structures of the pesticides studied in Chapters 2 and 5.............................. 26
Scheme 3.A.1. CB and o-DCB degradation routes achieved by Pseudomonas sp........... 71
Scheme 4.1. Suggested mechanism for the formation of 4-halo-1,2-dihydroxybenzene115
Scheme 4.2. Possible degradation pathways of p-halophenols, during TiO2 photocatalysis,
leading to either hydroquinone (HQ) (a and b) or benzoquinone (BQ) (d), and HQ
oxidation into BQ (c). For discussion, see text....................................................... 118
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1
CHAPTER 1.
Introduction and background of the thesis
1.1. Context
The total annual world production of synthetic organic chemicals is over 300 million tons
(Fewson, 1988) with more than 1000 new compounds fabricated every year, so that the
actual diversity of chemicals reaches approximately 20 millions of substances. For
instance, in Switzerland, 250 000 chemicals are commercialized.
If not treated, all these compounds sooner or later appear in the environment. In 1995,
Swoboda-Golberg (1995) surveyed the vast multitude of compound classes with
enormous number of functional groups and their possible combination and arrangements.
In this review, the major sources of synthetic organic chemicals have been summarized,
including industry, household and agriculture, with a special regard to compound
chemical stability or biodegradability.
In fact, during the past three decades, research on the impact of chemical pollution has
focused almost exclusively on the conventional ‘priority’ pollutants (i.e. persistent
organic pollutants, POPs) and this has been extensively reviewed recently (Birkett and
Lester, 2002). Today, these compounds are less relevant for many first world countries
because emissions have been substantially reduced through the adoption of appropriate
legal measures and the elimination of many of the dominant pollution sources. The focus
has consequently switched to compounds present in lower concentrations but which
nevertheless might have the ability to cause harm (Larsen et al., 2004). One of the
interesting characteristics of many of the chemicals that might cause this type of pollution
is that they do not need to be persistent in the environment to cause negative effects
(Ayscough et al., 2000). This is because their high transformation and removal rates can
be offset by their continuous introduction into the environment, often through sewage
treatment works (Suter and Giger, 2001).
2
In particular, volatile organic compounds (VOCs) of fuels and industrial solvents are
commonly found contaminants in the subsurface, pesticides and phenols in groundwater.
Persistent organic pollutants are long-lived organic compounds that become concentrated
as they move through the food chain. They are also known as persistent toxins that
bioaccumulate or as pseudo-oestrogenic chemicals. They have toxic effects on animal
reproduction, development, and immunological function. In a recent review, a special
attention has been focused on priority organic environmental contaminants and
experimental approaches for determination and studies of specific toxicity mechanisms
(Janošek et al., 2006).
In fact, although the risks associated with exposure to chemicals are probably most
significant with regard to the natural environment, the public’s concern is understandably
more focused on human exposure. However, the issue of xenobiotics (and their
metabolites) in the environment, notably the aquatic compartment, has been a growth
area in environmental chemistry for several years (Fent et al., 2006; Jones et al., 2001).
To date, most of the published literature has addressed the occurrence of organic
pollutants in sewage effluent and receiving waters. This is especially important for
human health in areas that practice indirect water reuse, where sewage effluent is released
to streams and rivers that are in turn used as a source of raw water for the production of
potable supplies for communities living downstream (Klöppfer, 1996a; Van Dijk-
Looijaard and Van Genderen, 2000). Unfortunately, there are extremely few data
available on the occurrence of pollutants in point-of-use drinking waters (that is, tap
water at the sink).
Of the Earth’s 1386 million cubic kilometers of water, only 2.5% of that quantity is
freshwater and nearly one third of this smaller amount is available for human use (Postel
et al., 1996). Against this backdrop of a finite water supply, total water withdrawn for
human uses has almost tripled in the last 50 years from 1382 km3 yr-1 in 1950 to 3973
km3 yr-1. Projections anticipate that the worldwide human water consumption will further
increase to 5235 km3 yr-1 by 2025 (Clarke and King, 2004): an amount greater than nine
times the annual flow of the Mississipi River (Goudie, 2000). Over 50% of the available
freshwater supplies is already used for human activities and this proportion is increasing
in response to growing agricultural, industrial, and residential demands (Postel et al.,
3
1996; Vorosmarty and Sahagian, 2000). By 2025 an estimate of 5 people out of 8 will be
living in conditions of water stress or scarcity (Arnell, 1999). Limiting water availability
will certainly alter ecosystems, human health, agricultural and industrial output, and
unfortunately increase the potential for conflict (Postel et al., 1996).
In the light of such alarming reports, two strategies are promising:
o developing and applying curative treatments is necessary to reduce the existing
pollution,
o adopting a prevention policy is a sustainable approach to preserve the overall
future life on Earth.
From the recent environmental awareness and legislative restrictions on uncontrolled
discharges of wastes, the best strategy for treatment of high toxic hazardous and
xenobiotic wastes may be their treatment at source with a special stress put on the
industry (Grüttner, 1997). A large group of aromatic compounds generally not produced
by natural processes has to be completely degraded and mineralized because of their high
persistence. This may be achieved through the development of particular solutions for
each ecological problem.
1.2. Conventional treatments for pollution removal
1.2.1. Biological treatments
1.2.1.1. Characteristics of biodegradation
It is clear that microbial-based treatment of contaminated environments and industrial
effluents are economically attractive in comparison with other conventional treatment
methods such as adsorption on granular activated carbon, air stripping or reverse osmosis,
combustion, and strong oxidative processes (Gogate and Pandit, 2004). In natural waters
and soils, the mineralization or complete biodegradation of an organic molecule is almost
always a consequence of microbial activity (Alexander, 1981). The most rapid and
complete degradation of the majority of pollutants is brought about under aerobic
conditions (Riser-Roberts, 1998).
4
The essential characteristics of aerobic microorganisms degrading organic pollutants are
illustrated in Figure 1.1. They consist of:
1. Metabolic processes for optimizing the contact between the microbial cells and
the organic pollutants. The chemicals must be accessible to the organisms having
biodegrading activities. For example, hydrocarbons are water-insoluble and their
degradation requires biosurfactants.
2. The initial intracellular attack of organic pollutants is an oxidative process, the
activation and incorporation of oxygen is the enzymatic key reaction catalyzed by
oxygenases and peroxidases.
3. Peripheral degradation pathways convert organic pollutants step by step into
intermediates of the central intermediary metabolism, e.g., the tricarboxylic acid
cycle.
4. Biosynthesis of cell biomass from the central precursor metabolites, e.g., acetyl-
CoA, succinate, pyruvate. Sugars required for various biosyntheses and growth
must be synthetized by gluconeogenesis.
A huge number of bacteria and fungia generally possess the capability to degrade organic
pollutants. Biodegradation is defined as the biologically catalyzed reduction in
complexity of chemical compounds (Alexander, 1994). It is based on two processes:
growth and cometabolism.
In the case of growth, organic pollutants are used as sole source of carbon and energy.
This process results in complete degradation (mineralization) of organic pollutants.
Cometabolism is defined as the metabolism of an organic compound in the presence of a
growth substrate which is used as the primary carbon and energy source.
5
Figure1.1. Main principle of aerobic biodegradation of hydrocarbons: growth associated
processes.
Enymatic key reactions of aerobic biodegradation are oxidations catalyzed by oxygenases
and peroxidases. Thus in the presence of molecular oxygen, compounds such as
benzoates, phenols, anilines and aromatic hydrocarbons are initially oxidized by mono-
or dioxygenases and converted to catechols or protocatechuates with subsequent
mineralization. The different pathways which are implicated to degrade xenobiotics have
been largely elucidated (Dagley, 1985). A special interest is now devoted to gene
expression for the regulation of biodegradative processes.
Fewson (1988) in his review presented the biodegradation process as environmentally
compatible; Singleton (1994) referred to “The Green Option”. In fact, biodegradation
may imply advantages as well as disadvantages for human beings and nature (Table 1.1).
Hydrocarbons
Initial attack by oxygenases
Degradation by peripheral pathways
Intermediatesof tricarboxylic
acid cycle
RespirationBios
ynthesi
s
Cell biomass
Growth
C4
C5
C6
C2
O2
CO2
O2
H2O
Fe 3+
SO42-
PO43-
NH4+
Hydrocarbons
Initial attack by oxygenases
Degradation by peripheral pathways
Intermediatesof tricarboxylic
acid cycle
RespirationBios
ynthesi
s
Cell biomass
Growth
C4
C5
C6
C2
O2
CO2
O2
H2O
Fe 3+
SO42-
PO43-
NH4+
6
Table 1.1. Beneficial and deleterious effects of biodegradation
The two first beneficial effects are quite well known, the third one only gained its
importance in recent years. It has been shown that microbial biotransformation could be
used to synthesize commercially useful products such as carboxylate diols and
fluorinated compounds (Ribbons et al., 1990). For example, Kiener (1995) reported a
process of microbial biotransformation of xylene and toluene via selective oxidation of
alkyl groups on aromatic N-heterocycles with following ring hydroxylation to
intermediates used for antipolytic and antidiabetic drug production.
1.2.1.2. Factors influencing microbial degradation of xenobiotics
Numerous factors can affect the biodegradation processes. They depend on the nature of
chemical molecules to be degraded (e.g. molecule size, charge, number and position of
functional groups, solubility and toxicity) as well as the environmental conditions. For
instance, some molecular features can increase recalcitrance (Fewson, 1988);
environmental factors influence the growth of organisms, the availability of xenobiotics
Beneficial effects Deleterious effects
Detoxification, decontamination and
bioremediation of hazardous or unpleasant
materials
Biodegradation of useful materials
Conversion to non-toxic products of
compounds that might give rise to harmful
substances
Production of toxic compounds from
innocuous precursors and conversion to
more recalcitrant forms
Formation of useful products Enhanced biodegradation of pesticides after
repeated applications
7
and more subtly, they can affect the gene expression (Van der Meer et al., 1998;
Daubaras and Chakrabarty, 1992).
Therefore, basic information is required to achieve a successful biotreatment:
o nature of the xenobiotic chemicals,
o concentration of the xenobiotic chemicals,
o presence and activity of the xenobiotic degrading microorganisms,
o appropriate conditions of cultivation which are required for the growth of the
target microorganisms.
Whenever they are identified and controllable, chemical and environmental factors
should be taken into account during the engineering development process and all along
the xenobiotic degradation (Liu et al., 2000).
Usual factors which play a significant role during biological treatments are:
o compound bioavailability,
o pH and temperature of the medium,
o nutrients,
o oxygen availability,
o residence time (dilution rate).
Laboratory experiments, which aim to study the xenobiotic biodegradation, are preferably
carried out at pH and temperature which are optimal for the cultivated microorganisms in
order to maximize their degradation activity. Many bacteria have their optimum pH near
neutrality and are mesophilic, whereas fungi prefer more acidic environments (Singleton,
1994). However, under suboptimal environmental conditions, an increased resistance of
microorganisms can be achieved by immobilization of the bacterial cells (Hörtemann and
Hempet, 1991). This is interesting for wastewater treatment.
Improving the monitoring of these parameters will turn out as an improved ability to
perform environmental risk assessments, to design environmental strategies and to
optimize wastewater treatment.
8
1.2.1.3. Biodegradability of xenobiotic compounds
Basically, biodegradability for a chemical compound refers to its potential for biological
degradation.
Biorecalcitrance for a chemical can (1) be due to the lack of the microbial enzymatic pool
necessary to degrade the molecule or (2) be attributed to its toxical properties against
microorganisms.
(1) In case of strains deficient in the adequate enzyme activities, biotreatment can
finally be achieved after involved acclimation and induction with the chemical (Haigler et
al., 1992; Spain and Gibson, 1988). This technique has been widely used for degradation
of halogenated aromatics (Reineke and Knackmuss, 1984; Spain and Nishino, 1987;
Seignez et al., 2001; Moos et al., 1983; Kle ka and Maier, 1985).
(2) In many situations, toxicity determination provides reliable information about
biodegradability of chemicals. This characteristic has been considered as a convenient
tool to predict the behavior of many substances in both environmental and treatment plant
situations (Mantis et al., 2005; Pessala et al., 2004; Chen et al., 1997; Boluda et al, 2002).
Several tests have been developed in order to assess the toxicity of chemical compounds
at different levels of ecotoxicity. For this purpose, many living organisms have been
selected for their relevance, accuracy and applicability. In particular, luminescent bacteria
bioassays (for eg., MicrotoxTM using Vibio fischeri) have emerged as a simple, fast and
comparatively inexpensive alternative to in-vivo bioassays with higher organisms (Parvez
et al., 2006; Steinberg et al., 1995; Manusad ianas et al., 2003; Kaiser and Palabrica,
1991).
Simultaneously, structure-activity relationships (QSAR) have been investigated in order
to assess toxicity and biodegradability, and therefore predict the potential ecological
effect of pollutants on various trophic levels (Schultz et al., 1988; Sixt et al., 1995;
Sarakinos et al., 2000; Chen et al., 1999). Various estimation programs even provide
qualitative and/or quantitative information concerning the biodegradability characteristics
of chemicals (Howard, 2000).
9
Considering the enormous quantity of scientific investigations related with microbial
degradation for diverse recalcitrant chemicals, it is presumable that every chemical are
biodegradable under specific conditions (presence or absence of oxygen; temperature;
pH; nutrient medium; liquid, gaseous or solid phases; microorganisms strains; etc.).
However in a context of bioremediation and/or pollution control (contaminated air, soils,
wastewaters, surface and groundwaters), biodegradability should mean biotreatability.
Indeed many methods which are aimed to assess biodegradability in fact prospect for
their possible removal in realistic conditions of treatment processes.
In chapter 2, the different methods for biodegradability assessment are proposed and
discussed.
In practice, it is frequent that no specific treatment methods are known so far for many
pollutants. In the worst cases their biological treatment is simply impossible to achieve.
In particular hardly-biodegradable anthropogenic substances accumulate and saturate the
auto-depurative conditions of the water cycle which becomes perturbed and overloaded.
The situation worsens because adequate water treatment systems are still missing or are
unable to lower the concentration of the toxic substances that represent a chronic
chemical risk (Klöpffer, 1996b, Andreadakis et al., 1997). In the meantime, numerous
water sources are deteriorated and compromise the overall supply and quality of drinking
water.
1.2.2. Conventional wastewater treatments
Water treatments have to focus on :
o contaminated drinking, ground and surface waters,
o wastewaters which contain toxic or biorecalcitrant compounds.
For this purpose, the actual and available technologies are very diverse, but conventional
processes are often classified as preliminary, primary, secondary and tertiary treatments
(Horan, 1990):
o During the preliminary step, debris and sandy materials are removed from the
wastewater.
10
o During primary treatment, wastewater is cleared of suspended solids (settlement)
and greases (floatation).
o Secondary treatment is a biological process: microorganisms absorb the dissolved
organic matter from wastewater and use them as food supply. Two strategies are possible
for aerobic wastewater treatment:
• the fixed biomass systems grow microorganisms on solid supports such as
rocks, sand or plastic. Wastewater is spread over the supports and by the way provides
organic matter and nutrients to the biomass: the biofilm grows by thickening. Trickling
filters, rotating biological contactors and sand filters are examples of fixed biomass
systems,
• suspended biomass systems maintain the biomass stirred during all the
treatment: microorganisms grow in size and number. After the treatment, the biomass is
settled out as sludge. Some of it is pumped back into the incoming wastewater to provide
seed microorganisms. The rest is sent to a further treatment process. Activated sludge,
extended aeration, oxidation ditch and sequential batch reactor systems are all examples
of suspended biomass systems.
o Tertiary treatment focuses on the removal of the disease-causing organisms from
wastewater. Treated wastewater can be disinfected by chlorine adjunction or ultraviolet
lightening. High levels of chlorine may be harmful to the aquatic life present in receiving
streams. Therefore treatment systems often add a neutralizing chemical to the treated
wastewater before stream discharge.
The incapability of conventional biological wastewater treatment to remove effectively
many industrial biorecalcitrant and/or toxic pollutants evidences that new efficient
treatment systems are needed. For the last 30 years, the water purification research has
been extensively growing. Rigorous pollution control and legislation in many countries
(for instance EEC, 1992; Conseil Fédéral Suisse, 1998) have resulted in an intensive
search for new and more efficient water treatment technologies.
11
1.2.3. Advanced Oxidation Processes (AOP)
In the last 15 years, a lot of research projects have been addressed to a special class of
oxidation techniques defined as Advanced Oxidation Process (AOP), pointing out its
potential prominent role in the wastewater purification (Ollis and Ekabi, 1993). It was
shown that AOP could successfully solve the problem of biorecalcitrant water pollutants
working at or near ambient temperature and pressure (Bahnemann et al., 1994; Bolton
and Cater, 1994).
All AOP are characterized by the same chemical feature: the production of hydroxyl
radicals OH• . These radicals are extremely reactive species and attack mainly every
organic molecules with rate constants usually in the order of 106-109 mol l-1 s-1 (Hoigne,
1997). OH• radicals are also characterized by a low selectivity of attack which is a
useful attribute for an oxidant used in wastewater treatment and for solving pollution
problems. The versatility of AOP is also enhanced by the fact that they offer different
possible ways for OH• radicals production, thus allowing a better compliance with the
specific treatment requirements.
Table 1.2 shows that OH• has the highest thermodynamic oxidation potential, which is
perhaps why OH• -based oxidation processes have gained the attention of many
advanced oxidation technology developers. In addition, most environmental contaminants
react 1 million to 1 billion times faster with OH• than with O3, a conventional oxidant
(US EPA, 1998).
Table 1.2 Oxidation potential of several oxidants in water (CRC Handbook, 1985)
Oxidant Oxidation potential (eV)
OH• 2.80
O(1D) 2.42
O3 2.07
H2O2 1.77
Permanganate ion 1.67
Chlorine 1.36
O2 1.23
12
The oxidation reactions involving hydroxyl radical and organic substrates (RH or PhX) in
aqueous solutions may be classified with respect to their character (Boossmann et al.,
1998):
o abstraction of hydrogen: OHRRHOH 2+•→+• (Eq 1.1)
o addition reactions: •→+• HOPhXPhXOH (Eq 1.2)
o electron transfer: [ ] −•+ +−→+• HOHRRHOH (Eq 1.3)
Application of AOP depends on the polluting load of wastes, as illustrated in Figure 1.2.
Figure 1.2. Suitability of water treatment technologies according to the Chemical Oxygen
Demand, COD (Andreozzi et al., 1999)
Only wastewater with relatively small Chemical Oxygen Demand (COD) contents ( 5 g
l-1) can be suitably treated by these techniques (Andreozzi et al., 1999) in order to avoid
an excessive consumption of expensive reactants. Wastes with massive pollutant contents
are more conveniently treated by wet oxidation or incineration (Mantzavinos et al, 1997;
Luck, 1999)
AOP can be classified according to the reaction phase (homogeneous or heterogeneous)
or to the methods used to generate the OH• radicals (chemical, electrochemical,
sonochemical or photochemical). The main used AOP are illustrated in Figure 1.3.
0 100010010
AOP
Wet oxidation
Incineration
Chemical Oxygen Demand, COD (mg l-1)
0 100010010
AOP
Wet oxidation
Incineration
Chemical Oxygen Demand, COD (mg l-1)
13
Figure 1.3. Main advanced oxidation processes (AOP).
An interesting class of AOP consists of the so-called Advanced Photochemical Oxidation
(APO) processes (US EPA, 1998). APO processes are characterized by a free radical
mechanism initiated by the interactions of photons of a proper energy level with the
molecules of chemical species present in the solution or with a catalyst.
The most used APO technologies can be broadly divided into the following groups: (1)
heterogeneous photocatalysis, (2) homogeneous photocatalysis, and (3) the photo-Fenton
process. These APO technologies are briefly described below.
1.2.3.1. Heterogeneous photocatalysis
Since 1972, when Fujishima and Honda (1972) reported the photocatalytic
decomposition of water on TiO2 electrodes, photocatalysis has been used with great
success in the degradation of a wide variety of contaminants, including alkanes, alcohols,
carboxylic acids, alkenes, phenols, dyes, PCBs, aromatic hydrocarbons, halogenated
alkanes and alkenes, surfactants, and pesticides (Ollis and Ekabi, 1993; Blake, 2001).
Heterogeneous photocatalysis is a technology based on the irradiation of a catalyst,
usually a semiconductor, which may be photoexcited to form electron-donor sites
(reducing sites) and electron-acceptor sites (oxidizing sites), providing great scope as
OH.
H2O2 / O3 / UV
UV / Fe3+ / H2O2
UV / TiO2 / H2O2
O3 / UV
H2O2 / O3 Sonolysis
Electrolysis
UV / TiO2
OH.OH.
H2O2 / O3 / UVH2O2 / O3 / UV
UV / Fe3+ / H2O2UV / Fe3+ / H2O2
UV / TiO2 / H2O2UV / TiO2 / H2O2
O3 / UVO3 / UV
H2O2 / O3H2O2 / O3 SonolysisSonolysis
ElectrolysisElectrolysis
UV / TiO2UV / TiO2
14
redox reagents. The process is heterogeneous because there are two active phases, solid
and liquid.
Among several semiconductors, titanium dioxide has proven to be the most suitable for
widespread environmental applications. TiO2 is stable to photo- and chemical corrosion,
and inexpensive. TiO2 has an appropriate energetic separation between its valence and
conduction bands (VB and CB, respectively), which can be surpassed by the energy of a
solar photon. The VB and CB energies of the TiO2 are estimated to be +3.1 and -0.1
volts, respectively, which means that its band gap energy is 3.2 eV and absorbs in the UV
light (wavelength λ<387 nm). Hydroxyl radical ( OH• ) is the main oxidizing species
responsible for photo-oxidation of the organic compounds (Legrini et al., 1993;
Valavanidis et al., 2006). The first event, after absorption of near ultraviolet radiations at
λ<387 nm, is the generation of electron/hole pairs (Eq 1.4) separated between the CB and
VB.
( )+− +→+ heh 22 TiOTiO ν (Eq 1.4)
Some of the numerous events which take place after the UV light absorption by TiO2
particles are presented in figure 1.4. The subsequent generation and separation of
electrons ( −e CB) and holes ( +h VB) are also mentioned. In particular, three oxidation
reactions have been experimentally observed: electron transfer from R , OH 2 and OH−
adsorbed on the catalyst surface.
It has been shown that the photocatalytic degradation of a water contaminant can be
enhanced by addition of H2O2, a better electron acceptor than O2 (Ollis et al., 1991;
Pichat et al., 1995; Diller et al., 1996). This effect has been explained by the formation of
hydroxyl radicals during the reaction with CB electrons by Eq 1.5:
-OHOHOH 22 +•→+ −CBe (Eq 1.5)
It limits the electron-hole recombination rate and increases the hydroxyl radical
concentration at the TiO2 surface.
15
Figure 1.4. Schematic of an irradiated TiO2 semi-conductor particle with some possible
photo-chemical and photo-physical processes.
1.2.3.2. Homogeneous photodegradation
The former applications of homogeneous photodegradation (single-phase system) to treat
contaminated waters concerned the use of UV/H2O2 and UV/O3. The use of UV light for
photodegradation of pollutants can be classified into two principal areas:
o direct photodegradation, which proceeds from direct excitation of the pollutant by
UV light, and
o photo-oxidation, where light leads oxidative processes principally initiated by
hydroxyl radicals. This process involves the use of an oxidant to generate radicals, which
attack the organic pollutants to initiate oxidation. The three major oxidants used are:
hydrogen peroxide (H2O2), ozone, and photo-Fenton system (Fe3+/H2O2).
UV
Energy
h+
e-
VB
CB
TiO2
O2
O2
H2 O/ -OH ; R
OH ; R +
Photo-reduction
Photo-oxidation
-
e - + H2 O2-OH + OH
OH + R CO2 + H 2 O
Exc
itat
ion
Rec
ombi
nati
on
intermediates
OH
UV
Energy
h+
e-
VB
CB
TiO2
O2
O2
H2 O/ -OH ; R
OH ; R +
Photo-reduction
Photo-oxidation
-
e - + H2 O2-OH + OHe - + H2 O2-OH + OH
OH + R CO2 + H 2 O
Exc
itat
ion
Rec
ombi
nati
on
intermediates
OH
16
1.2.3.3. Fenton and photo-Fenton assisted reactions
The Fenton reaction was discovered by H.J.H. Fenton in 1894. In 1934, Haber and
Waiser postulated that the effective oxidative agent in the Fenton reaction was the
hydroxyl radical ( OH• ). Since then, some groups have tried to explain the whole
mechanism (Walling, 1975; Prousek, 1995; Sychev and Isak, 1995).
The Fenton reaction can be outlined as follows:
OHOHMOHM nn •++→+ −+++ )1(22 (Eq 1.6)
where M is a transition metal such as Fe or Cu.
In absence of light and complexing ligands other than water, the most accepted
mechanism of H2O2 decomposition which occurs in acid homogeneous aqueous solution,
involves the formation of hydroxyperoxyl ( •2HO / −2O ) and hydroxyl radicals ( OH• ).
(DeLaat and Gallard, 1999; Gallard and De Laat, 2000)
The OH• radical, once in solution, attacks almost every organic compound. The metal
regeneration can follow different pathways. For Fe2+, the most accepted scheme is
described in Eq 1.7 to Eq 1.14 (Sychev and Isak, 1995).
OHOHFeOHFe •++→+ −++ 322
2 K= 63 l mol-1 s-1 (Eq 1.7)
+++ +•+→+ HHOFeOHFe 22
223 K= 3.1 x 10-3 l mol-1 s-1 (Eq 1.8)
−++ +→•+ OHFeOHFe 32 K= 3 x 108 l mol-1 s-1 (Eq 1.9)
OHHOOHOH 2222 +•→+• K= 3.3 x 107 l mol-1 s-1 (Eq 1.10)
22
23 OHFeHOFe ++→•+ +++ K= 2 x 103 l mol-1 s-1 (Eq 1.11)
22
23 OFeOFe +→•+ +−+ K= 5 x 107 l mol-1 s-1 (Eq 1.13)
−++ +→•+ 23
22 HOFeHOFe K= 1.2 x 106 l mol-1 s-1 (Eq 1.14)
17
Fenton reaction rates are strongly increased by irradiation with UV/visible light (Rupert
et al., 1993; Sun and Pignatello, 1993; Bandara et al., 1996). This type of photoassisted
reaction is referred to as the “photo-Fenton reaction” (Zepp et al., 1992).
+++ +•+→++ HOHFehOHFe 22
3 ν (Eq 1.15)
The positive effect of irradiation on the degradation rate is due to the photochemical
regeneration of ferrous ions ( +2Fe ) by photoreduction of ferric ions ( +3Fe ) (Faust and
Hoigne, 1990). The new generated ferrous ion reacts with H2O2 producing a second
OH• radical and ferric ion (Eq 1.7), and the cycle continues. In these conditions, iron
can be considered as a real catalyst.
The main advantage of the photo-Fenton process is the light sensitivity up to a
wavelength of 600 nm (35% of the solar radiation) (Safarzadehamiri et al., 1997). The
depth of light penetration is high and the contact between pollutant and oxidizing agent is
close, since homogeneous solution is used (Bauer et al., 1999; Fallmann et al., 1999).
Fenton and photo-Fenton processes have been applied with great success to treat a wide
variety of contaminants, including phenols (Kiwi et al, 1994; Gernjak et al., 2003), dyes
(Bandara et al., 1996), halogenated alkanes and alkenes (Buyuksonmez et al., 1999), and
pesticides (Fallmann et al., 1999; Parra et al., 2000).
Several parameters governing or influencing the kinetics of the photo-Fenton system
have been studied: pH (Scott et al., 1995), iron concentration (Krutzler and Bauer, 1999),
iron species (Cuzzola et al., 2002), H2O2 concentration (Safarzade-Hamiri et al., 1997),
initial pollutant concentration (Rupert et al., 1993), temperature (Sagawe et al., 2001;
Rodriguez et al., 2002), irradiation source (Rodriguez et al., 2002), presence of radical
scavengers (Chen and Pignatello, 1997; Kiwi et al., 2000).
18
1.3. Aim of the thesis
The aim of this thesis is to propose a strategy to treat xenobiotic compounds. As
introduced in part 1.2., the Advanced Oxidation Processes (AOP) appear as interesting
tools, in comparison with the well-established techniques like activated carbon, air
stripping, reverse osmosis, combustion, etc. Indeed, many of these techniques only
transfer the pollutants from one phase to another without destroying them. Other
processes are selective but rather slow in degradation rates, or rapid but not selective,
thus increasing the engineering, maintenance and energy costs. Other restrictive factors
are economic reasons, oxidative potential, effluent characteristics or tendency to form
harmful disinfection by-products as, for example, the case of formation of
trihalomethanes (THMs) when a chlorination procedure is used for drinking water
treatment.
Biological treatment is limited to wastewaters which contain biodegradable substances
and which are not inhibitive nor toxic to the bioculture.
Although AOP are cheaper than combustion or wet oxidation technologies, the most
serious drawback of AOP is their relatively high operational costs, compared to those of
biological treatments.
For these reasons, their use as a pre-treatment step for the enhancement of the
biodegradability of wastewater containing recalcitrant or inhibitory compounds can be
justified when the intermediates resulting from the AOP are readily degradable by
microorganisms. Thus considering AOP as a preliminary treatment before inexpensive
biological process seems very promising from an economical point of view (Robertson,
1996; Scott and Ollis, 1995; Scott and Ollis, 1997; Sarria et al., 2002).
Therefore, the approach proposed in this thesis consists of:
1. Determining the potential of a target xenobiotic to be degraded through a
biological process,
2. Depending on the previous response, applying the appropriate treatment among
the three most promising processes described before: biodegradation,
photocatalysis or integrated biological-photocatalytic treatment.
19
Figure 1.5 presents the schematic of this strategy.
Figure 1.5. Overview of the strategy proposed to treat xenobiotics
Each situation will be illustrated in this thesis. For this purpose, the manuscript is
organized in six chapters. In the first chapter, background and fundamentals of the
research are presented. The second chapter clarifies the assessment methods used to
determine the biodegradability of chemicals. Chapters 3, 4 and 5 are devoted to
biodegradation, photocatalytic process and the integrated biological-photocatalytic
treatment, respectively. These methods are applied to specific xenobiotic substances.
Each definite topic is subject to publication. By the way, chapters 2 to 5 provide the
information through separated scientific articles. The following paragraph 1.4 details the
several target xenobiotics which have been studied in this work. It is also aimed to
introduce the different aspects related to both the substances and their appropriate
degradation methods.
XenobioticsXenobiotics
?Biodegradability
?No
AdvancedOxidationProcesses
AOP
PHOTOCATALYSIS
TiO2, photo-Fenton
Chemistry / fundamentals
SARS : Structure-Activity Relationships
Yes
BiologicalTreatments
BIOENGINEERING
Microbiological topicsProcess enhancement
Limited
IntegratedAOP / Biological
Processes
BIOENGINEERING
Feasibility
Microbiological topics
Process enhancement
ChapterChapter 44
ChapterChapter 55
ChapterChapter 22
ChapterChapter 33
p-halophenols
pesticides
VOCs(Volatile Organic
Compounds )
XenobioticsXenobiotics
?Biodegradability
?No
AdvancedOxidationProcesses
AOP
PHOTOCATALYSIS
TiO2, photo-Fenton
Chemistry / fundamentals
SARS : Structure-Activity Relationships
Yes
BiologicalTreatments
BIOENGINEERING
Microbiological topicsProcess enhancement
Limited
IntegratedAOP / Biological
Processes
BIOENGINEERING
Feasibility
Microbiological topics
Process enhancement
ChapterChapter 44
ChapterChapter 55
ChapterChapter 22
ChapterChapter 33
p-halophenols
pesticides
VOCs(Volatile Organic
Compounds )
20
1.4. The target xenobiotics compounds
1.4.1. Volatile Organic Compounds
It is worth mentioning that there is no agreement about the definition of Volatile Organic
Compounds (VOCs). In spoken language, VOCs are often used as synonyms for organic
solvents. An effect-oriented definition, mainly used in the USA, states that VOCs are all
organic compounds contributing to photochemical ozone creation. More general
definitions are based on physical and chemical properties of the compounds, such as
chemical structure, boiling point, air/water partitioning, and vapor pressure.
Definitions based on vapor pressure are frequently used. In the USA, VOCs are defined
as organic compounds that have a vapor pressure more than 13.3 Pa at 25°C, according to
ASTN test method D3960–90. In the European Union, a common definition is that VOCs
are organic compounds with a vapor pressure above 10 Pa at 20°C (European VOC
Solvents Directive 1999/13/EC). The Australian National Pollutant Inventory defines a
VOCs as chemical compounds based on carbon chains or rings (and also containing
hydrogen) with a vapor pressure greater than 2 mm of mercury (0.27kPa) at 25°C,
excluding methane.
VOCs are a topic of interest in many disciplines, such as food, flavor, fragrance and
medical sciences. The main area dealing with VOCs may be the environmental
chemistry, because VOCs contribute to stratospheric ozone depletion, tropospheric ozone
formation, toxic and carcinogenic human health effects, etc.
Chlorobenzenes and TEX compounds (toluene, ethylbenzene and xylenes) are among the
most important contaminants present in surface and groundwater which usually originate
from the leakage of tanks containing petroleum-derived products and industrial
wastewaters (Akunar-Askar et al., 2003). Because they are both toxic and relatively water
soluble compared with other petroleum constituents, their entry into surface and drinking
water is of major concern.
Many technologies were developed and optimized for removing either chlorinated
benzenes or TEX compounds from industrial effluents and groundwater in order to limit
the contamination of downstream drinking water resources. Among all remediation
21
technologies for treating TEX-contaminated groundwater, bioremediation appears to be
an economical, energy-efficient and environmentally sound approach (Devinny et al.,
1999; Shim et al., 2002; Nishino et al., 1992). Microorganisms are able to degrade
chlorinated benzenes and TEX under aerobic, microaerobic or hypoxic, as well as
anaerobic conditions. However these substances are highly toxic and easily provoke
inhibitions to the microorganisms which degrade them (Sardessai and Bhosle, 2002).
Therefore developing and upgrading biotechniques for simultaneous and efficient
removal of all TEX remains challenging (Rula and Cohen, 1999; Shim and Yang, 1999;
León et al., 1999).
In this thesis, the target VOCs consisted of chlorobenzene, o-dichlorobenzene, toluene,
ethylbenzene and o-, m-, p-xylenes (TEX). Their molecular structure is illustrated in
Scheme 1.1. Two other chlorinated solvents (dichloroethane, dichloromethane) are
occasionally mentioned since they are studied for specific topics.
Scheme 1.1. Structures of the VOCs studied in Chapters 2 and 3.
CH3
C2H5
CH3
CH3
CH3
CH3
CH3
CH3
TOLUENE
XYLENES (o, m, p)
ETHYLBENZENE
Cl CHLOROBENZENE
Cl
Cl
o-DICHLOROBENZENE
22
In chapter 2, these substances will be shown biodegradable. As biological treatment
prevails over the other mentioned-processes, it is hence proposed to study and enhance
their biodegradation. The basis of the work dealing with VOCs is three-fold:
1. Develop and test a reliable engineering process with the aim to produce a bacterial
consortium able to degrade chlorobenzenes and/or TEX,
2. Characterize and optimize the degradation process, focusing on microbial, kinetic and
biochemical aspects,
3. Automate the cultivation procedure of the efficient chlorobenzenes/TEX-degrading
bacterial consortium, using LabVIEW© software.
This topic is presented in chapter 3.
1.4.2. Phenols
Phenols, another group of organic compounds, are considered as priority pollutants since
they are harmful to organisms at low concentrations and many of them have been
classified as hazardous pollutants because of their potential to harm human health. It
should be noted that the contamination of drinking water by phenolics, at even a
concentration of 0.005 mg l-1 could bring about significant taste and odor problems
making it unfit for use. Human consumption of phenol-contaminated water can cause
severe pains leading to damage of capillaries ultimately causing death. Phenol containing
water, when chlorinated during disinfection of water also results in the formation of
chlorophenols. The most important pollution sources containing phenols and phenolic
compounds such as nitrophenols, chlorophenols and other halophenols, are the
wastewaters from the iron-steel, coke, petroleum, pesticide, paint, solvent, pharmaceutics,
wood preserving chemicals, and paper and pulp industries (Kahru et al., 1998a).
Current methods for removing phenolics from wastewater include microbial degradation,
adsorption on activated carbon, chemical oxidation (using agents such as ozone,
hydrogen peroxide or chlorine dioxine), deep-well injection, incineration solvent
extraction and irradiation. Solvent extraction methods are expensive and deep-well
injection may lead to contamination of groundwater. Adsorption and oxidation treatments
23
become exceedingly expensive when low effluent concentrations must be achieved.
Wastewaters containing phenol in the range of 5-500 mg l-1 are considered suitable for
treatment by biological processes. A wide variety of pure and mixed cultures of
microorganisms are even capable of degrading phenols and phenolic compounds under
both aerobic and anaerobic conditions. Surprisingly, although biological treatment has
shown great promises, microbial degradation of these compounds is seen as a cost
effective method. Indeed biological treatment of phenolics is limited by the intrinsic
properties of these compounds owing to their toxicity; they are slow to biodegrade and
the degrading microorganisms must be exposed to only low concentrations of the
substrates (Kahru et al., 1998b; Grau and Darin, 1997).
This is the reason why alternative technologies have to be explored (Bandara et al., 2001;
Espuglas et al., 2002; Ghaly et al., 2001; Tang and Huang, 1996). In this context, and
according to the strategy proposed in this thesis, AOP was selected for treatment of
phenolic substances.
In chapter 4, TiO2 photocatalysis is considered for the removal of p-halophenols (see
Scheme 1.2) in contaminated water, with the aim to better understand the chemical
mechanisms involved during their photocatalytic degradation. For that purpose, structure-
activity relationships and molecular descriptors are employed to characterize
photodegradability of p-halophenols. Toxicological properties are also investigated.
Phenol is used as reference compound in order to assess the implication of halogen in the
photodegradation of p-halophenols.
Scheme 1.2. Structures of the phenolic substances studied in Chapters 2 and 4.
X HO p-HALOPHENOL (X= Cl, F, Br, I)
PHENOL HO
24
1.4.3. Pesticides
Advancing increase of production and application of pesticides for agriculture as well as
for plant protection and animal health has caused the pollution of soil, ground and surface
water which involves a serious risk to the environment and also to the human health due
to direct exposure or through residues in food and drinking water (Ma et al., 2005; Ma et
al., 2002).
In the world, alarming levels of pesticides have been reported to be persistent, toxic,
mutagenic, carcinogenic, and tumorogenic (Costa, 2005). Pesticide contamination of
water systems has been of major concern in recent years. Pesticide residues reach the
aquatic environment through manufacturing plants, direct surface run-off, leaching,
careless disposal of empty containers, equipment washings, etc. (Canna-Michaelidou and
Nicolaou, 1996).
Pesticides are divided into 3 families depending on the living organisms that they target.
They can also be classified according to their molecular structure (organochlorine,
organophosphorus, phenyl urea, etc.). Due to the wide range of pesticides it is extremely
difficult to apply one single method for pesticide removal.
Chemical treatments can be inefficient for the removal of synthetic organochlorine
pesticides from waters and may also lead to hazardous final products (Drzewicz et al.,
2004; Chiron et al., 2000). On the other hand, advanced water treatment processes, and
mainly the adsorption onto activated carbon, have proved to be efficient and reliable
methods for removal of aqueous-dissolved organic pesticides. The main drawbacks of
these techniques are their cost and the fact that substances remain untouched.
In recent years, the ability of microorganisms to metabolize some pesticides has also
received much attention due to the environmental persistence and toxicity of these
chemicals. Although in some cases, microbial metabolism of contaminants may produce
toxic metabolites, a variety of microorganisms (mainly aerobic bacteria and fungi) are
known to utilize organic pesticides as the sole carbon or energy source, but in specific
conditions (Perrin-Ganier et al., 1996). Practically conventional activated sludge systems
often fail to achieve high efficiency in removing pesticides from wastewater. It is mainly
25
due to the low biodegradability of organic pesticides, their toxicity or their tendency to
inhibit microorganisms ( etkauskaité et al., 1998).
Besides the Advanced Oxidation Processes, and particularly the photocatalytic processes,
were shown particularly efficient for degradation of the organic pollutants which are
recalcitrant to conventional treatment because of their high chemical stability and/or low
biodegradability (Gogate et al., 2004a; Konstantinou et al., 2003; Pera-Titus et al., 2004;
Rodriguez et al., 2002; Sarria et al., 2002).
Due to their relatively high operational costs compared to biotreatments, their use as a
pre-treatment step has been considered as a solution to enhance biodegradability of
problematic wastewater. Such a technology is justified if the intermediates resulting from
the degradation of the toxic or recalcitrant compounds are readily degraded by
microorganisms. Therefore, combinations of AOP as preliminary treatments with
inexpensive biological processes become very promising from an economical and
environmental point of view (Pulgarin et al., 1999; Parra et al., 2002; Pulgarin and Kiwi,
1996).
Several water-soluble pesticides were selected: Alachlor, Atrazine, Chlorfenvinphos,
Diuron, Isoproturon and Pentachlorophenol (of special interest because of their extremely
easy transport in the environment, seriously threatening all surface and groundwater).
These substances are included among the Priority Substances (PS) listed by the European
Commission (OJEC, 2001).
Their molecular structure is illustrated in Scheme 1.3.
In chapter 2, the above-mentioned pesticides are tested for their biodegradability.
As their biotreatability is shown difficult to achieve, the third case proposed by the
strategy (Figure 1.5), i.e. the integrated photocatalytic-biological treatment, is applied to
the pesticides.
Chapter 5 is devoted to present these experiments.
26
Scheme 1.3. Structures of the pesticides studied in Chapters 2 and 5.
The basis of the work dealing with pesticides is three-fold:
1. Assess the feasibility of the integrated photocatalytic-biological technology for the
removal of the single or mixed pesticides using several analytical tools and testing
methods,
2. Improve the modelling relation between chemical, biological and toxicological
parameters,
3. Optimize the coupling strategy, that is to say select the best moment for coupling.
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37
CHAPTER 2.
Biodegradability Assessment of Several Priority Hazardous Substances: Choice, Application and Relevance Regarding Toxicity and Bacterial Activity
2.1. Abstract
Nineteen compounds listed in the category of Priority Substances (PS) were selected for a
biodegradation study using standardized tests. The compounds consist of pesticides,
chlorinated solvents and Volatile Organic Compounds (VOCs). In this chapter, the choice
of the most suitable method is discussed in relation to the physico-chemical properties of
each substance. Zahn-Wellens, Manometric Respirometry and Closed-Bottle tests are
alternatively used. Experimental results are presented and interpreted. Toxicity
(MicrotoxTM) and bacterial viability (Bac-lightTM) are also used as tools to investigate the
influence of each substance on the microbial population (activated sludge). In addition,
experimental values are compared with predictive data calculated according to
Quantitative Structure Activity Relationships (QSARs) models. Biodeg Models were
permitted to correctly estimate 17 substances; Survey Models and screening tests also
revealed the same behavior for 16 target compounds.
2.2. Introduction
Chemical release into environment is one of the major patterns of the anthropogenic
pollution. Xenobiotics tend to persist in the natural cycles unless they are degraded.
Therefore developing efficient treatment systems enables to reduce pollution, since these
systems possibly lead to innocuous substances. Considering both natural and laboratory-
controlled scales, bacteria are the most significant tool involved in ultimate
Accepted for publication in Chemosphere
38
biodegradation (Howard et al., 1975; Alexander, 1994). For that topic, mixed microbial
culture prevail over pure cultures. Indeed even if pure cultures are useful in identifying
possible pathways of degradation, organisms isolated by enrichment may not necessarily
be active in the environment.
Biodegradability prospect of a target substance has to meet three main ecological points
(Pitter and Chudoba, 1990): Is bacterial adaptation necessary? What are the rate and the
degree of biodegradation under given conditions? Is it possible to reach the ultimate
mineralization or are toxic compounds produced? To answer these questions, many
testing strategies have been evolved. In 1990, Pitter and Chudoba proposed a
classification in accordance with the Organization of Economic Cooperation and
Development (OECD, Paris).
Three groups of tests were defined:
o ready biodegradability (or screening),
o inherent biodegradability, and
o simulation.
Screening tests indicate if a compound is degradable under natural conditions without any
problem. Six methods permit the screening of chemicals for ready biodegradability in an
aerobic aqueous medium: Dissolved Organic Carbon (DOC) Die-Away, Modified OECD
Screening -DOC Die-Away, Carbon dioxide (CO2) Evolution (Modified Sturm Test),
Manometric Respirometry, Closed Bottle and MITI (Ministry of International Trade and
Industry -Japan). These informative tests basically discriminate readily biodegradable
compounds from others, but they often underestimate the potentiality of degradation in
environmental systems.
Therefore when the result is negative, inherent (potential) biodegradability tests are
required. In this category of tests, the concentrations of both the target compound and
inoculum (usually activated sludge) are higher than with the screening tests, and the ratio
is shifted in favor of microorganisms which are adaptable for long periods.
Inherent biodegradability tests provide information concerning the biotraitability of a
substance. For example, the Zahn-Wellens procedure is an interesting test to prospect the
39
“behavior” of a chemical in an activated sludge treatment plant, since the experimental
conditions are similar to this process.
Finally, simulation test methods confirm the positive or negative results of the previous
tests.
Reproducibility and reliability of all the tests were improved consequently with the use of
standardized protocols. Indeed, some test guidelines were developed by various national
and international organizations such as OECD, U.S. Environmental Protection Agency
(EPA) Office of Pollution Prevention and Toxics (OPPT), and the European Commission.
They list experimental conditions, analytical methods, and criteria for whether a chemical
is considered to be biodegradable (pass) and non-biodegradable (fail) (Howard et al.,
1987).
Nevertheless sometimes it is difficult to obtain precise informative data concerning the
biodegradability of pesticides or volatile organic compounds (VOCs) (Cowan et al.,
1996). Literature still provides heterogeneous results, mainly due to various initial
concentrations of target chemicals or to the many studies devoted to natural or industrial
matters. However a methodology for translating OECD tests to degradation rates into
realistic environmental situations has been presented, but only applied to 20 chemicals
(Struijs and Van den Berg, 1995) or a limited number of substances (Federle et al., 1997).
Since the 1980s, many studies have prevailed to build up estimation programs which able
to provide qualitative and/or quantitative information concerning the biodegradability
characteristics of chemicals (Howard, 2000). They aimed to bypass experimental
limitations and to better understand the implemented mechanisms. However, probably
due to their novelty, only a few external validations have been completed until now
(Boethling et al., 2003).
In this chapter, first, a procedure for selecting the most suitable test of biodegradability
regarding the physico-chemical properties of 19 xenobiotics and considering a processing
in an industrial wastewater treatment plant (WWTP) is discussed. Then the adequate
methods for the testing of the biotraitability of the target compounds are presented.
Finally the results of the experimental biodegradability testing are commented and
40
compared with some estimates which are calculated with the most relevant predicting
methods according to the review completed by Howard (2000).
The 19 xenobiotics consist of seven medium/high soluble pesticides (Alachlor, Atrazine,
Chlorfenvinphos, Diuron, Isoproturon, Lindane and Pentachlorophenol) listed as Priority
Substances (PS) according to the European Commission (OJEC, 2001). The others are
chlorinated solvents (Dichloroethane, Dichloromethane), Volatile Organic Compounds,
VOCs (Chlorobenzene, o-Dichlorobenzene, Toluene, Ethylbenzene and o-, m-, p-
Xylenes). Finally three other pharmaceutical substances were also tested (“T824”,
C11H11Cl2N2O, (±)-alpha-(2,4-dichlorophenyl)-1H-imidazole-1-ethanol; “R113675”;
C17H21NO3; (-)-[4aS-(4a alpha, 6 beta, 8aR*)]- 4a, 5, 9, 10, 11, 12- hexahydro- 3-
methoxy- 11- methyl- 6H-benzofuro [3a,3,2-elf] [2] benzazepin- 6-ol hydrobromide;
“FEMAM”, C9H11NO2, alpha-methyl-alpha-amino-phenyl-acetinamide).
2.3. Material and methods
2.3.1. Chemicals
The pesticides Alachlor (purity: 95%), Chlorfenvinphos (93.2%) and Diuron (98.5%)
were supplied by Aragonesas Agro (Spain); Atrazine (95%) and Isoproturon (98%) by
Ciba (Switzerland); Lindane (90%) by Exagama 90 Rhône-Poulenc (France) and
Pentachlorophenol (98%) by Aldrich (Switzerland). Dichloromethane (99.8%),
Dichloroethane (99.5%), Chlorobenzene (98%) and o-Dichlorobenzene (>98%) were
from Merck (Switzerland). Toluene (99.5%) and Ethylbenzene (99.8%) were provided by
Acros Organics (Belgium) and o-, m- and p- Xylenes (>98%) were from Fluka
(Switzerland). Other chemicals C11H10Cl2N2O “T824”, C17H21NO3 “R113675” and α-
methylphenylglycynamide “Femam” were provided by pharmaceutical industries. Their
physico-chemical properties are presented in Table 2.1. Glucose (97%) was from Sigma
(Switzerland). Diethylene glycol C4H10O3 (>98%) used as reference compound in the
Zahn-Wellens test came from Fluka (Switzerland).
41
Table 2.1. Physico-chemical properties of the target compounds (Tomlin, 1997; Boublick
and Cervenkova, 1984; SRC, Hansch and Leo, 1995). Patm = 1 atm = 760 mmHg
Compound Molecular
formula
Molecular
weight,
g/mol
Solubility
in water at
25°C,
mg/l
Vapor pressure,
mm Hg at 25 °C
Octanol-water
partition
coefficient,
log Kow
Alachlor C14H20ClNO2 269.77 239 2200 10-8 3.52
Atrazine C8H14ClN5 215.69 34 28.9 10-8 2.61
Chlorfenvinphos C12H14Cl3O4P 359.58 125 750 10-8 3.81
Chlorobenzene C6H5Cl 112.56 490 12.00 2.84
o-Dichlorobenzene C6H4Cl2 147.00 150 1.50 3.38
Dichloroethane C4H8Cl4 197.92 5101 - 1.83
Dichloromethane CH2Cl2 84.93 13001 435.00 1.25
Diuron C9H10Cl2N2O 233.10 41 6.9 10-8 2.68
Ethylbenzene C8H10 106.17 160 9.75 3.15
FEMAM C9H12N2O 164.00 94300
Isoproturon C12H18N2O 203.29 65 2.47 10-8 2.87
Lindane C6H6Cl6 290.83 8 4200 10-8 3.72
Pentachlorophenol C6HCl5O 266.34 15 11000 10-8 5.12
R113675 C17H21NO3 368.27 31000 1.09
Toluene C7H8 92.14 530 28.50 2.73
T824 C11H10Cl2N2O 257.12 131 2.67
o-Xylene C8H10 106.17 170 6.75 3.12
m-Xylene C8H10 106.17 160 8.25 3.20
p-Xylene C8H10 106.17 180 9.00 3.15
42
2.3.2. Biodegradability methods
2.3.2.1. Zahn-Wellens test
The Zahn-Wellens test was carried out according to the EC protocol (Directive
88/302/EEC). The inoculum –activated sludge- came from a secondary effluent of the
treatment plant of Morges (Switzerland). The fresh sample was centrifuged at 10 000
RPM during 7 minutes at 20 °C, and washed once with mineral medium.
The official protocol of the Zahn-Wellens test recommends that compound
concentrations fit with the limits of the sensibility of the Dissolved Organic Carbon, DOC
determination (50 mgc/l) and also with the hydrosolubility thresholds of each substance
(see Table 2.1).
According to the guidelines of the Zahn-Wellens test, the ratio between the carbon
content of the tested substance and the dry-weight of the inoculum was ranged between 1
and 4 (experimental values averaged 3.3).
Aeration and homogenization were guaranteed.
Preliminary experiments were realized to check that neither volatilization nor adsorption
occurred during the testing period.
2.3.2.2. Closed bottle and Manometric respirometry tests
The Closed bottle test (Directive 92/69/EEC), also called BOD28 – Biochemical Oxygen
Demand, measured during 28 days – and the Manometric respirometry test (Directive
92/69/EEC) were executed using a Hg free WTW 2000 Oxitop© unit thermostated at 20
°C.
Experiments were carried out according to official protocols, except for the composition
of the mineral medium (the same as for the Zahn-Wellens test) and the use of the Oxitop©
material.
This equipment facilitated the manipulation and minimized the quantity of vessels
required. Indeed when O2 is consumed, the CO2 produced is trapped by sodium
hydroxide pellets and a depression in the bottle is automatically recorded.
43
The volume of the solutions poured in the bottles was determined according to the
theoretical oxygen demand without nitrification (ThODNH3). The formula is:
( )M
ThODNH
o- 2/na 2/5p 3s 2/3n)-cl-(h 2c 163
++++= ,
with Cc Hh Clcl Nn Nana Oo Pp Ss, the formula of the test compound, and M, its molecular
weight.
The average concentration of biomass was 25 mg of suspended solids per liter of mineral
nutrient medium.
2.3.3. Initial concentrations
All the target compound concentrations were fixed to 50 mg/l, except for the FEMAM
(500 mg/l) or whenever the hydrosolubility threshold was lower (Lindane 2 mg/l,
Atrazine 20 mg/l, Pentachlorophenol 15 mg/l). These concentrations corresponded to
carbon contents higher than the minimal threshold prescribed by the official guidelines
for the biodegradability testing.
2.3.4. Analytical methods
Pesticide concentration was analysed using reverse-phase liquid chromatography (flow 1
ml/min) with UV detector in a HPLC-UV (Varian 9012, 9100, 9065) with an ODS-2
column (Waters 4.6x250 mm, from Phenomenex) and a guard column (Waters 4.6x10
mm): Alachlor (H2O/Acetonitrile 40/60, 224nm), Atrazine, Isoproturon, Chlorfenvinphos
and Diuron (Acetic acid(1%)/Acetonitrile 80/20-40/60 (0-30 min) and 40/60 (30-35 min),
249 nm).
Concerning Pentachlorophenol, the HPLC system was constituted by a LC-10 AT VP
pump (Shimadzu) and a UV–Visible diode array detector (Agilent 1100 Series). The
stationary phase was a Hypersil ODS Teknokroma column (250x4.6 mm). A
methanol/water mixture (80/20) was used as the mobile phase, which was degassed by
sonication and filtered (0.45 µm) before using.
44
The VOCs concentrations were determined by using a gas chromatograph (Varian Star
3400 Cx, Palo Alto, USA), equipped with a flame ionization detector and a DB-624
capillary column (30m x0,53mm) (J&W Scientific, Folsom, USA).
The Dissolved Organic Carbon, DOC was measured using a Total Organic Carbon, TOC
analyser (Shimadzu model 5050A, Japan) calibrated with standard solutions of potassium
hydrogen phthalate, sodium carbonate and sodium hydrogen carbonate. Limit of
detection of the apparatus was 0.2 mgC/l. The mineralization degree (degradation percent)
corresponded to the ratio of DOC disappearance between sample and blank.
The pH was determined using a combined electrode and pH Meter from Radiometer
Copenhagen PHM 92 (Denmark).
The toxicity was assessed using the MicrotoxTM test system. This test was carried out
using a Microtox model 500 analyzer (Canada); this laboratory-based temperature
controlled photometer (15–27 °C) maintains luminescent bacteria reagent (Vibrio
fischeri) and test samples at the appropriate test temperature. This self-calibrating
instrument measures the light production from the luminescent bacteria reagent. The
sample toxicity is determined by measuring the effective concentration from which 50%
of the light is lost due to compound toxicity (EC50).
The Toxic Unit (TU) is used to express the data, as proposed by U.S. EPA (Lankford and
Eckenfelder, 1990): TU = 100/EC50.
The LIVE/DEAD® BacLightTM Bacterial Viability Kit L-7007 (Molecular Probes
Corporation, Canada) provided a quantitative index of bacterial viability. In fact it uses a
mixture of SYTO® 9 green fluorescent nucleic acid stain which labels bacteria with both
intact and damaged membranes. Besides, the red fluorescent nucleic acid stain Propidium
Iodide penetrates only bacteria with damaged membranes. The bacterial samples are
centrifuged and filtered through 0.2 µm PC Memb 13 mm (Nucleopore CORP., Canada).
The Epi-FluoroMicroscopic observation is carried out under UV light with emission
filters BP 450-490 nm, reflector FT 510 and stop-filter LP 520 (Eclipse 800, Nikon,
Switzerland).
In all cases the 28 day-tests were duplicated and gave similar results.
45
For each experimental sample, the chemical analyses were replicated thrice. Standard
deviation was less than 5%.
In this study, reported results correspond to the average value of the measurements.
2.4. Results and discussion
2.4.1. Selecting the appropriate test of biodegradability
In this study biodegradability of 19 chemicals is prospected to assess their biotraitability
in an industrial WWTP. Considering this objective, the Zahn-Wellens test is the most
adequate procedure, because the experimental conditions are close to a WWTP process.
Unfortunately this method is inappropriate for many compounds due to their physico-
chemical properties. Indeed this method is only applicable to water-soluble and non-
volatile substances.
Concerning hydrosolubility, threshold is prescribed by the minimal carbon content of the
tested solution. Since 50 mg of carbon per liter are requisite, Atrazine, Diuron, Lindane
and Pentachlorophenol are out of range (Table 2.1).
Besides VOCs and many target pesticides tend to volatilize. More precisely without any
value recommended in the Zahn-Wellens guidelines, compounds which present a vapor
pressure included between the atmospheric pressure (Patm = 760 mm Hg) and 7.5 10-2 mm
Hg (threshold for VOC) can be considered as moderately volatile (Alachlor, Atrazine,
Chlorfenvinphos, Lindane, Pentachlorophenol).
Therefore the use of the Zahn-Wellens protocol turns out to be inconsistent for these
substances, as well as for VOCs.
Consequently two adapted Ready-Biodegradability methods were chosen to test them.
Actually target compounds are supposed biorecalcitrant or at least difficult to treat (they
are defined as hazardous substances (OJEC, 2001)). So Ready-Tests experiments are
likely to fail. However as previously stated, negative results obtained with Ready-Tests
do not prove unbiotraitability, whereas positive results guarantee the feasibility of a
biotreatment in an industrial WWTP.
46
Table 2.2. Test methods and results for the biodegradability assessment of the target
compounds. Degradation percent is related with chromatographic results (1), DOC
measurements (2) or theoretical oxygen uptake (3).
Test
Methods Compounds
Degradation
(%)
Time
(d)
Initial
Concentration
(mg/l)
Conclusion
Chlorfenvinphos 60 (1+2)
30 (1+2)
22
28
10-50
125 Inhibitive factor
Diuron <0 (2) 28 50 Cellular lyses
FEMAM 16 (1+2) 28 500
Isoproturon 0 (1+2) 28 50
R113675 94 (2) 6 50 Biodegradation
Zahn-Wellens
T824 <0 (2) 56 50 Cellular lyses,
enzymatic lacks, no bacterial adaptation
Alachlor 0 (1+3) 28 50 Inhibitive factor
Chlorfenvinphos 0 (1+3) 28 50 Cellular lyses,
chlorine toxicity
Dichloroethane 25 (1+3) 26 50 Microbial adaptation
Manometric Respirometry
Dichloromethane 0 (1+3) 28 50 Cellular lyses
Atrazine 100 (1+3) 21 20 Biodegradation
Chlorobenzene 73 (1+3) 20 50 Biodegradation
o-Dichlorobenzene <0 (1+3) 28 50 No cellular damage
Ethylbenzene 81 (1+3) 14 50 Biodegradation
Lindane 100 (1+3) 9 2 Biodegradation
Pentachlorophenol 0 (1+3) 28 15
Toluene 93 (1+3) 14 50 Biodegradation
m-, p-Xylene 85 (1+3) 14 50 Biodegradation
BOD28
o-Xylene 80 (1+3) 19 50 Biodegradation
47
Therefore a compromise is aimed between the more selective experimental conditions of
the Ready-Tests and the unsuitable conditions of the Zahn-Wellens test.
To approximate to the Zahn-Wellens test conditions, two modifications are considered:
o a slight concentration of the test compound (due to its low hydrosolubility threshold)
o a closed system (in order to avoid volatilization).
In fact the slightly adapted screening tests applied are the Closed-Bottle Test (BOD28)
and the Manometric Respirometry. They are favored because both are operated in closed
vessels.
In particular, the Closed-Bottle test lengthened up to 28 days is favorably disposed
towards low hydrosolubility values.
Table 2.2 summarizes the different methods applied for each compound.
2.4.2. Zahn-Wellens test
The Zahn-Wellens method was used to test the biodegradability of Diuron, Isoproturon,
and the three pharmaceutical compounds (R113675, T824 and FEMAM).
During 28 days, the contents of DOC were measured in both samples and blanks (only
sludge in mineral medium). As indicated in the protocol, the mineralization degree,
expressed as the “degradation percent”, was calculated regularly. It corresponds to the
ratio of DOC disappearance between sample and blank, taking into account the initial
carbon content in both sample and blank. In fact samples were performed every two or
three days.
Values indicated in Figure 2.1 correspond to average values calculated from at least six
experimental data. Standard deviation and experimental error were less than 0.05 mgc/l.
They are not pictured in order to clarify the experimental results.
Diethylene Glycol (DEG) was used as a reference substance to check the correct
functioning of experimental system. As shown in Figure 1, DEG was degraded up to 70%
in ten days, which validates the initial activity of the activated sludge used in the Zahn-
Wellens test. One of the pharmaceutical compounds, R113675 was also degraded up to
94% after six days, with a constant pH. Therefore this substance is considered as
48
potentially biodegradable. On the contrary, no DOC was removed in 28 days for Diuron,
Isoproturon, FEMAM and T824.
Figure 2.1. Zahn-Wellens biodegradability test. Evolution of the degradation percent as a
function of time for R113675, T824, FEMAM, diuron, isoproturon and diethylene glycol
DEG (control). Glucose adding is symbolized with the dark circles. Data are related with
DOC analyses and correspond to the average value of at least six measurements.
First of all, Isoproturon biorecalcitrance confirms the conclusions obtained previously
with the same test (Parra et al., 2002).
Regarding Diuron, this pesticide is considered as non-ready biodegradable, according to
MITI database (www.cerij.or.jp/ceri_en/otoiawase/otoiawase_menu.html). Nevertheless
many studies have been completed to improve biodegradation carried out by specific
bacteria or fungi, such as Aeromonas sp., Arthrobacter sp. and other commercially
available strains (Tixier et al., 2000; Tixier et al., 2002).
Precise pathways have even been described (Turnbull et al., 2001) and many material
supports have been tested: soil, natural waters, etc. (Sumpono et al., 2003). It means that
microbial degradation can be performed with specific and acclimated strains and
reinforces the need of biodegradability studies in order to assess the potential risks of
-40
-20
0
20
40
60
80
100
0 7 14 21 28Time (days)
Deg
rada
tion
(%)
Diuron
DEG-control
FEMAM
Isoproturon
R113675
T824
Glucoseadding
49
substances in the environment (Liu et al., 2000). In fact cellular lysis is suspected for
bacteria unacclimated to Diuron and T824, because basification and carbon release were
recorded during the experiment.
Furthermore, microscopic observations revealed that almost 80% of the bacterial
population showed damaged membranes on day 28 with the BacLightTM assay. This
statement needs to be distinguished from an apparent cellular damage due to starvation.
Indeed it is likely that the bacteria have long ago used up any available nutrients, and
have in fact been starving for a long time. But blanks (bacteria with mineral medium)
revealed that only 20% of the total population was affected by this factor. So 60% of the
damaged cells remained assigned to the presence of hazardous substances.
As presented in Figure 2.1, the consumption of glucose added after 28 days of experiment
reveals that many cells able to degrade this easily biodegradable substrate were still
viable.
Biorecalcitrance of T824 and Diuron in the tested conditions should more result from
their structural stability towards microbiological attack rather than from the toxicity of
the solution.
This statement was concomitant with the evolution of the overall toxicity of the tested
solutions. For example, the MicrotoxTM assay carried out with T824-samples which were
harvested on days 0 and 28 revealed that toxicity (1/EC50) decreased from 14.5 to 0.4 TU
in the course of the biodegradability testing (results not shown).
2.4.3. “Ready-biodegradability” tests
Ready-biodegradability testing was used to assess the biotraitability of 14 xenobiotics
using two modified screening methods: BOD28 and Manometric Respirometry (Table
2.2).
The 14 target compounds evaluated with the tests fall into four classes. Degradation
could be inexistent, complete or partial; the last case was shown to be favoured by
bacterial adaptation or not.
Figure 2a indicates the evolution of the degradation percent for Dichloroethane and
Dichloromethane (Manometric Respirometry) and for VOCs (BOD28) during 28 days.
Figure 2b illustrates the evolution of the degradation percent during the 28 days of
50
experiment, i.e. Manometric Respirometry for Alachlor and Chlorfenvinphos, and BOD28
for Pentachlorophenol, Atrazine and Lindane.
In both tests the degradation percent was evaluated according to the compound
concentration (chromatographic analyses) and to the oxygen consumption due to the
biodegradation of the target compound. Indeed oxygen consumption was measured in the
closed vessels containing either compound with sludge, compound without sludge or
sludge without compound. As for the Zahn-Wellens test (Figure 2.1), Figures 2.2a and
2.2b indicate average values in order to simplify them.
Figure 2.2a. Evolution of the VOC degradation percent during 28 days for the Ready-
Biodegradability testing (BOD28). Data are related with chromatographic results and
correspond to the average value of at least six measurements.
First of all, VOCs illustrated in Figure 2.2a (Toluene, Ethylbenzene, o-, m- and p-
Xylenes) passed over 70% of degradation in approximately one week. So they can be
considered as biodegradable. Chlorobenzene was also degraded within 3 weeks. Both
stripping and adsorption were avoided.
As shown in Figure 2a, Dichloroethane was degraded up to 25% in 26 days (GC
analyses). This partial elimination performed in drastic conditions for biodegradation
-20
0
20
40
60
80
100
0 7 14 21 28Time (days)
Deg
rada
tion
(%)
Chlorobenzene
Dichlorobenzene
Dichloroethane
Dichloromethane
Ethylbenzenze ; m- and p-Xylenes
Toluene
o-Xylene
51
testing indicates a possible biodegradation in a WWTP. A faster process would certainly
be achieved with a pre-adapted biomass. This finding is in consistency with a reported
bio-oxidative activity of wastewater microbiota under a specific incubation method,
which led to the same range of degradation (Tabak et al., 1981).
On the contrary, no degradation of Dichloromethane was observed after 28 days of
experiment (chromatographic analyses and oxygen depletion measurements). During the
testing of Dichlorobenzene, carbon release has been observed, due to cellular lyses.
Figure 2.2b. Evolution of the degradation percent during 28 days for the Ready-
Biodegradability testing of Alachlor, Chlorfenvinphos (Manometric Respirometry),
Atrazine, Lindane and Pentachlorophenol (BOD28). Data are related with chromatographic
results and correspond to the average value of at least six measurements.
Concerning Pentachlorophenol and Alachlor (Figure 2.2b), neither oxygen consumption
nor disappearance of the initial compound (HPLC analysis) were observed during 28
days. Thus, no degradation occurred and bacteria appeared to be inhibited, as glucose
remained untouched when added after the test.
Alachlor was previously shown to inhibit the biodegradation ability of mixed bacterial
cultures up to 18% at a 50 mg/l concentration (Zargoc-Koncan, 1996); and
Pentachlorophenol is considered relatively resistant to biodegradation due to its high
-20
0
20
40
60
80
100
0 7 14 21 28Time (days)
Deg
rada
tion
(%)
Atrazine
Chlorfenvinphos
Lindane
Pentachlorophenol,Alachlor
52
chlorine content. Incidentally its biodegradation can be performed with pre-adapted
bacteria (Arthrobacter sp., for eg.), but remains limited by the substrate inhibition
(Klecka and Maier, 1985; Liu et al., 2000).
As illustrated in Figure 2.2b, Atrazine was degraded up to 100% within 21 days. This
tendency is supported by previous results (Zagorc-Koncan, 1996). Particularly they
showed that despite its harmful effect on algae, Atrazine (0-70 mg/l) caused no effect on
biodegradation processes occurring with mixed bacterial populations.
The low hydrosolubility of Lindane (see Table 2.1 and Figure 2.2b) imposes a maximal
concentration of 2 mg/l, which corresponds to an oxygen demand of 1,3 mg/l. Despite
this minimal value, a depletion of oxygen was recorded and maintained after 7 days of
experiment, taking into account blanks containing either biomass without Lindane or
compound without inoculation.
The aerobic inactivity observed in the controls suggests a bacterial involvement.
Furthermore, the recovered sludge was able to consume glucose, proving its state of
health.
Kipopoulou et al. (2004) confirmed the occurrence of the combined sorption and
biodegradation processes on sludge, which is different to photolysis, hydrolysis and air
stripping (Mills, 1985; Dorussen and Wassenberg, 1997). They also strengthened that
biosorption was the main removal process involved in the environment and that it was
linked with the bioconcentration of Lindane.
Concerning Chlorfenvinphos, no degradation was observed after 28 days of experiment
(Figure 2b). Despite the measured-20% losses due to volatilization, biodegradation has
occurred up to 60% within 22 days for 10 to 50 mg/l of Chlorfenvinphos, and only up to
30% for 125 mg/l (results not shown). Each bacterial culture exhibited a first step of
cellular lyses detected by microscopic observations (damaged cells), basification of the
cultivation medium and carbon release. The lower the ratio compound/bacteria, the
further the biodegradation occurred. So it seems that Chlorfenvinphos implies irreversible
cellular damages, which can be overwhelmed if the bacterial population is concentrated
enough.
Experimental conclusions are summarized in Table 2.2.
53
2.4.4. Predictive estimation models
Some predictive models can provide screening-level estimates of biodegradability. Their
benefits consist in their large applicability, their relevance and their obvious technical
advantages. Their development also results from many studies devoted to link the
chemical structure of a compound with its biodegradability (Mani et al., 1991).
Howard (2000) reviewed the most reliable estimation methods. In this study, five of them
were applied to each of the 19 target compounds and the conclusions were matched to the
experimental data obtained with the Zahn-Wellens and Ready-Biodegradability tests
stated above.
2.4.4.1. Qualitative Substructure Model
First, the Qualitative Substructure Model, QSM (Niemi et al., 1987) focuses on the
presence of some specific structural sub-units which are associated with predicting half-
lives (T1/2).
This method was built on the basis of BOD data converted into T1/2. For eg., a chemical
with a theoretical BOD of 50% or greater in five days was assumed to have a T1/2 of five
days and a 25% theoretical BOD in ten days corresponded to a T1/2 of 20 days.
Using these data, specific structural units exhibited by persistent chemicals were
identified and classified.
In this study, each target compound was divided into its molecular sub-units. Each of
them was associated with the corresponding T1/2 as indicated in the QSM. This analysis
revealed that every target compound had at least one persistent molecular segment.
As indicated in Table 2.3, QSM overestimated T1/2 for only one compound, R113675.
In fact, the estimated T1/2 greater than 100 days was due to the highly branched molecular
structure of R113675, but experimentally, the Zahn-Wellens test showed that the
compound was biodegradable in approximately six days (Table 2.2).
Conversely the presence of a “benzene ring with at least two substitutions (non-hydroxyl)
and a log Kow >2.18” (Niemi et al., 1987) appeared to be a crucial molecular
characteristic for many of the studied compounds. This characteristic corresponds to a
54
half-life longer than 100 days in the QSM and indeed each involved molecule was
experimented biorecalcitrant with our screening-tests (Alachlor, Chlorfenvinphos,
Diuron, o-DCB, Isoproturon, T824 and Pentachlorophenol, see Table 2.2).
So it can be concluded that experimental data agreed with QSM predictions for 18
xenobiotics.
2.4.4.2. Biodegradability Probability Program
The second type of predictive methods concerns the Biodegradability Probability
Program (Howard et al., 1992; Boethling et al., 1994), which requires more calculation.
In these four models, some coefficients are attributed to specific structural fragments.
Then they are introduced into mathematical formulas.
Biodeg Models, also named BioWIN 1 and 2 (Boethling et al., 2003) provide coefficients
which were developed by linear (BW1) and non-linear (BW2) regressions. They forecast
a “fast” or a “not fast biodegradation”.
Concerning the Survey Models, also named BioWIN 3 and 4, they distinguish primary
biodegradation (BW3) from ultimate biodegradation (BW4).
In this study, fragment coefficients were calculated separately for each target compound
according to the proposed mathematical equations and taking into account both types and
quantities of specific molecular fragments. Then a probability of biodegradation was
calculated, depending on molecular weight, empirical equation constants and fragment
coefficients.
All detailed information concerning the specific fragments and their regression-derived
coefficients are provided in the corresponding model descriptions (Boethling et al., 2003;
Howard, 2000).
Table 2.3 summarizes the results obtained for the biodegradability assessment of the 19
xenobiotics using the four BioWIN models (noted BW 1-4), the Qualitative Structural
Model (noted QSM) and the Zahn-Wellens and screening tests (noted Exp).
55
Table 2.3. Comparison of experimental results and estimated data for the biodegradability
assessment of the 19 xenobiotics (Exp:Experimental, NB:NonBiodegradable,
B:Biodegradable, QSM:Qualitative Structural Model (T1/2 indicated in days),
BW:BioWIN, F:Fast, NF:Not Fast, D:Days, W:Weeks, M:Months)
Biodegradability Assessment Methods Compounds
Exp QSM (T1/2) BW 1 BW 2 BW 3 BW 4
Conclusion
Alachlor NB >100 d NF NF W M +
Atrazine B >100 d NF NF W M -
Chlorfenvinphos NB >100 d NF F W M +/-
Chlorobenzene B >5 d F NF D W +/-
o-Dichlorobenzene NB >100 d NF NF W M +
Dichloroethane NB >15 d NF NF W M +
Dichloromethane NB >15 d NF NF W M +
Diuron NB >100 d NF NF W M +
Ethylbenzene B 2-15 d F F DW W +
FEMAM NB >15 d F F W W -
Isoproturon NB >100 d F F W M +/-
Lindane B >40 d NF NF W M -
Pentachlorophenol NB >100 d NF NF W M +
R113675 B >100 d F F W M +/-
Toluene B 15 d F F D W +
T824 NB >100 d NF NF W M +
m-, p-Xylene B 2-15 d F F D W +
o-Xylene B 2-15 d F F D W +
56
First, if each compound is considered separately, the comparison between the five
estimates shows that the five predictive methods give coherent results for every target
compounds, except for Chlorfenvinphos, Chlorobenzenze and R113675.
In fact, Chlorfenvinphos was expected to biodegrade fast according to the sole BW2,
because this model attributed an excessive contribution to the phosphate ester group. In
this case, both experiments and other models invalidated this prediction.
Conversely, BW2 attributed a slow biodegradability to Chlorobenzene because of the
presence of the aromatic chloride.
Concerning R113675, it was due to the highly branched molecular structure which had
already been noticed by the QSM. However, as for Chlorfenvinphos, the other results
obtained for Chlorobenzene and R113675 were opposite.
Thus the five predictive models provided coherent results for 16 xenobiotics.
Secondly, the comparison between experiments and estimates shows that experiments
agreed with predictions for 14 xenobiotics.
In fact, Atrazine and Lindane were surprisingly experimented biodegradable whereas
they were predicted resistant. Their ability to be degraded by a bacterial community was
discussed previously. However since the predictions were pessimistic, no environmental
problem would have resulted from this inaccuracy.
On the contrary, FEMAM and Isoproturon were determined degradable according to
models but were experimented biorecalcitrant with the Zahn-Wellens test.
For these two substances, BW1 and BW2 performed poorly.
Concerning Isoproturon, BW4 and QSM predicted that mineralization would be achieved
in three months at least, and so the compound was presented rather resistant, which was
not the case for Femam. The remark concerning BW1 and BW2 was already noticed by
Rorije et al. (1999). The authors also explained that both consistency and overall
accuracy could be improved by optimizing the classification criterion for these two
models using the MITI data (Tunkel et al., 2000).
57
Finally the overall results suggest that the QSM and Survey Models (BW3 and 4) are
generally suitable for use in biodegradability assessment, regarding a consecutive
industrial WWTP process.
2.5. Conclusion
In this chapter, the most suitable procedures for testing biodegradability of 19 xenobiotics
were selected regarding the physico-chemical properties of the compounds and
considering a consecutive processing in an industrial wastewater treatment plant
(WWTP). The adequate methods have been successfully performed in order to test the
substance biotraitability.
In this study experimental results agreed with predictions for 16 xenobiotics, whatever
their biodegradability may be.
The usefulness of estimation models was shown to be considerably increased whenever
several methods are performed comparatively.
This conclusion reinforces the strategy suggested by Boethling et al. (2003) who
proposed to allow prediction only when several models give unanimous results.
By this approach, screening level experiments may be avoided with limited
environmental risks.
2.6. References
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USA.
Boethling, R.S., Howard, P.H., Meylan, W., Stiteler, W., Beauman, J., Tirado, N., 1994.
Group contribution method for predicting probability and rate of aerobic
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Boethling, R.S., Lynch, D.G., Thom, G.C., 2003. Predicting ready biodegradability of
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Boublik, T., Cervenkova, I., 1984. Vapor pressure, refractive indexes and densities at
20.0 degree C, and vapor-liquid equilibrium at 101.325 kPa in the tert-amyl methyl
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Cowan, C.E., Federle, T.W., Larson, R.J., Feijtel, T.C., 1996. Impact of biodegradation
test methods on the development and applicability of biodegradation QSARs. SAR
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Dorussen, H.L., Wassenberg, W.B., 1997. Feasability of treatment of low polluted
wastewater in municipal wastewater treatment plants. Wat. Sci. Technol. 35, 73-78.
Federle, T.W., Gasior, S.D., Nuck, B.A., 1997. Extrapolating mineralization rates from
the ready CO2 screening test to activated sludge, river-water and soil. Environ.
Toxicol. Chem. 16, 122-134.
Hansch, C., Leo, A., 1995. Exploring QSAR. American Chemical Society, Washington,
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Howard, P.H., 2000. Biodegradation. In: Mackay, D. (Ed.). Handbook of Property
Estimation Methods for Chemicals. Environmental and Health Sciences. Lewis, Boca
Raton, FL, USA, pp. 281-310.
Howard, P.H., Boethling, R.S., Stiteler, W., Meylan, A.E., Hueber, A.E., Beauman, J.,
Larosche, M.E., 1992. Predictive model for aerobic biodegradability developed from
a file of evaluated biodegradation data. Environ. Toxicol. Chem. 11, 593-603.
Howard, P.H., Hueber, A.E., Boethling, R.S., 1987. Biodegradation data evaluation for
structure/biodegradation relations. Environ. Toxicol. Chem. 6, 1-10.
Howard, P.H., Saxena, P.R., Durkin, P.R., Ou, L.T., 1975. Review and evaluation of
available techniques for determining persistence and routes of degradation of
chemical substances in the environment. EPA 560/575/006. Final report. U.S.
Environmental Protection Agency, Washington, DC.
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Kipopoulou, A.M., Zouboulis, A., Samara, C., Kouimtzis, T., 2004. The fate of lindane in
the conventional activated sludge treatment process. Chemosphere 55, 81-91
Klecka, G.M., Maier, W.J., 1985. Kinetics of microbial growth on pentachlorophenol.
Appl. Environ. Microbiol. 49, 46-53.
Lankford, P., Eckenfelder Jr., W., 1990. Toxicity Reduction in Industrial Effluents. Van
Vostrand Reinhold, New York, USA.
Liu, D., Maguire, R.J., Lau, Y.L., Pacepavicius, G.J., Okamura, H., Aoyama, I., 2000.
Factors affecting chemical biodegradation. Environ. Toxicol. 15, 476-483
Mills, W.B., 1985. Water quality assessment: a screening procedure for toxic and
conventional pollutants in surface and groundwater-Part I. EPA 600/685/002a. U.S.
Environmental Protection Agency, Washington, DC
Mani, S.V., Connell, D.W., Braddock, R.D., 1991. Structure activity relationships for the
prediction of biodegradabiltiy of environmental pollutants. Crit. Rev. Enviro. Contr.
21, 217-236
Niemi, G.J., Veith, R.R., Regal, R.R., Vaishnav, D.D., 1987. Structural features
associated with degradable and persistent chemicals. Environ. Toxicol. Chem. 6, 515-
527
OJEC (Official Journal of the European Communities). 15.12.2001. L331/1. Decision N°
2455/2001/EC.
Parra, S., Malato, S., Pulgarin, C., 2002. New integrated photocatalytic-biological flow
system using supported TiO2 and fixed bacteria for the mineralization of isoproturon.
Applied Catalysis B: Environmental 36, 131–144.
Pitter P., Chudoba J., 1990. Biodegradability of Organic Substances in the Aquatic
Environment. CRC Press, Boca Raton, FL, USA, pp. 306.
Rorije, E., Loonen, H., Muller, M., Klopman, G., Peijnenburg, W., 1999. Evaluation and
application of models for the prediction of ready biodegradability in the MITI-I test.
Chemosphere 38, 1409-1417.
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Struijs, J., Van den Berg, R., 1995. Standardized biodegradability tests: Extrapolation to
aerobic environments. Water Res. 29, 255-262.
Sumpono, Perotti, P., Belan, A., Forestier, C., Lavedrine, B., Bohatier, J., 2003. Effect of
Diuron on aquatic bacteria in laboratory-scale wastewater treatment ponds with
special reference to Aeromonas species studied by colony hybridization.
Chemosphere 50, 445-455.
Syracuse Research Corporation, SRC. http://www.syrres.com.
Tabak, H.H., Quave, S.A., Mashni, C.I., Barth, E.F., 1981. Biodegradability studies with
organic priority pollutant compounds. J. Water Pollut. Control Fed. 53, 1503-1518
Tixier, C., Bogaerts, P., Sancelme, M., Bonnemoy, F., Twagilimana, L., Cuer, A.,
Bohatier, J., Veschambre, H., 2000. Fungal biodegradation of a phenylurea herbicide,
Diuron: structure and toxicity of metabolites. Pest. Manag. Sci. 56, 455-462.
Tixier, C., Sancelme, M., Ait-Aissa, S., Widehem, P., Bonnemoy, F., Cuer, A., Truffaut,
N., Veschambre, H., 2002. Biotransformation of phenylurea herbicides by a soil
bacterial strain, Arthrobacter sp. N2: structure, ecotoxicity and fate of diuron
metabolite with soil fungi. Chemosphere 46, 519-26.
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Tunkel, J., Howard, P.H., Boethling, R.S., Stiteler, W., Loonen, H., 2000. Predicting
ready biodegradability in the japanese Ministry of International Trade and Industry
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61
CHAPTER 3.
Biodegradation of Volatile Organic Compounds
PART A.
ENHANCING PRODUCTION OF ADAPTED BACTERIA TO DEGRADE CHLORINATED
AROMATICS
3.A.1. Abstract
The study was aimed to enhance the production of adapted bacteria for degrading
chlorobenzene (CB) and o-dichlorobenzene (DCB). Batch cultures fed by pulses of
substrate (SPB, Substrate Pulse Batch process) were successfully performed in a 12 liter-
bioreactor: 240 mg l-1 h-1 of CB/DCB mixture was eliminated and 50 mg l-1 h-1 of dry
weight produced. Studying the substrate removal dynamics, concentration of metabolic
intermediates, chlorides and cellular integrity allowed the correct proceeding of bacterial
cultivation. The substrate loading rate was optimized to obtain the highest biomass
productivity with the lowest toxicity and inhibiting effects. The time interval between
each pulse was decreased gradually so that a continuous substrate feeding was finally
attained. In the continuous feeding technique, both cell productivity and CB/DCB
elimination capacity were twice the values obtained with the SPB. Results come up to
industrial prospects and present the continuous feeding technique as an attractive
adaptation of SPB to degrade chlorinated solvents.
Accepted for publication in Industrial & Engineering Chemistry Research
62
3.A.2. Introduction
At present, the purification of industrial waste air containing volatile organic compounds
(VOCs) increasingly refers to microbiological methods, due to the rapid development of
industrial biotechnology. The use of biofilters is based on the microbial degradation of
organic substrates (xenobiotics) to carbon dioxide and water. Contrary to chemical
methods (catalytic or high-temperature burning), adsorption or filtration through selective
membranes which only transfer the pollution, the microbial mineralization allows a
sustainable pollution control of the environment. Biological techniques for waste gases
purification are operated onto a great variety of organic compounds, including
chlorinated substances. Indeed chlorinated benzenes are widely used as organic solvents,
heat transfer agents and insecticides, and their production as intermediates in the
synthesis of chemicals, such as pesticides and dyes also contributed to spread out these
compounds into the environment (Pearson, 1982). In particular Chlorobenzene (CB) and
1, 2-dichlorobenzene (DCB) were identified as priority pollutants by the US
Environmental Protection Agency (Oh and Bartha, 1994).
The biomass responsible for the degradation of the gaseous contaminant fed to the
biofilter can be supplied by the support material itself (Barona et al., 2004). Nevertheless,
the filter-bed inoculation with active microorganisms is a frequent and necessary
prerequisite for successful operation (Kennes and Thalasso, 1998; Kennes and Veiga,
2001). Consequently, prior to the inoculation itself, the proper selection, acclimation and
activation of the microorganisms is crucial to ensure the successful operation of the
biofilters. Although pure strains of microorganisms can be purchased, mixed cultures can
be obtained from the sludge originated in wastewater treatment plants or similar origins
(Prado et al., 2003; Prado et al, 2005). Once the biomass is in the laboratory, the best
process to grow specific strains was shown to be batch or microcosms cultures (León et
al, 1999).
This study was aimed to adopt Substrate Pulse Batch (SPB) technique for the efficient
biomass production for the biodegradation of chlorobenzene (CB) and o-
dichlorobenzenze (DCB). SPB technique is a batch cultivation process in which
substrates are supplied in pulses. This technique can be considered as a special case of a
63
conventional, semi-continuous extended culture operation: total volume remains virtually
constant because substrate is fed in a gaseous, solid or very concentrated liquid form,
according to an intermittent profile.
In this chapter, the substrate loading rate in SPB technique was optimized to obtain the
highest biomass productivity with the lowest toxicity and inhibiting effects. For this
purpose the time interval between each pulse was decreased gradually so that a
continuous substrate feeding was finally attained. Finally, the results of SPB technique
were compared with that of continuous process in terms of CB/DCB elimination capacity
and biomass productivity.
3.A.3. Materials and methods
3.A.3.1. Inoculum Preparation
The bacterial consortium used to inoculate the bioreactor was selected in our laboratory
in a previous work (Seignez et al, 2001). The bacteria were originally issued from the
sludge of a wastewater treatment plant (Novartis and Rohner AG, Basel, Switzerland)
and from the biotrickling filters treating chlorinated aromatic solvents and preserved at -
80°C (Rohner AG, Basel, Switzerland).
A sample of this adapted mixed culture was cultivated in flasks containing nutrient
medium (CM67, Oxoid LTD, Basingtoke, England) at 30°C for 1 day.
Then the obtained biomass was collected by centrifugation and transferred into the batch
reactor.
The nutrient medium consisted of inorganic salts and vitamins necessary for the growth
of microorganisms, the same composition as previously reported with ammonium as the
nitrogen source (León et al, 1999). The medium was composed of: 3.3 g l-1 K2HPO4, 1.9
g l-1 NaH2PO4.2H2O, 4.5 g l-1 (NH4)2SO4, 0.2 g l-1 MgSO4.7H2O, 0.5 g l-1 CaCl2.2H2O, 10
ml l-1trace-elements and 2.5 ml l-1 vitamins. The trace-element solution contained 5 g l-1
CaCl2.2H2O, 1 g l-1 EDTA.2H2O, 1 g l-1 FeSO4.7H2O, 160 g l-1 MnCl2.4H2O, 40 g l-1
ZnSO4.7H2O, 30 g l-1 H3BO3, 40 g l-1 CoCl2.2H2O, 40 g l-1 CuCl2.2H2O, 4.6 g l-1
NiCl2.H2O and 40 mg l-1 NaMoO4.2H2O. The vitamin solution consisted of 8 g l-1 biotin,
64
100 mg l-1 p-aminobenzoic acid, 0.2 g l-1 nicotinic acid, 0.2 g l-1 thiamin-Cl, 0.2 g l-1 Ca-
pantothenate, 0.2 g l-1 pyridoxine-HCl, 20 mg l-1 cyanocobalamin, 80 mg l-1 riboflavine,
40 mg l-1 folic acid, 0.2 g l-1 choline-Cl and 0.8 g l-1 myo-inositol.
The sole carbon and energy source consisted in 80.8 % (m/m) chlorobenzene (CB) and
19.2 % o-dichlorobenzene (DCB). The volumetric ratio of 4:1 corresponds to industrial
gas effluents emission of two CBs from Rohner SA chemical industry (Basel,
Switzerland). Compounds were supplied by Fluka, Buchs, Switzerland. Experiments
were made under non-sterile conditions. The volatile organic carbons (VOCs) i.e., CB
and o-DCB were supplied by a precision feeding pump (SpectraPhysics, San José, Calif.).
3.A.3.2. Bioreactor and Automation Program
The experimental set-up used in this study is presented in Figure 3.A.1.
Bioreactor is made of stainless steel, with a total volume of 20 l. The working liquid
volume was 12 l. Numerous ports along its height and at the bottom allow feeding,
sampling and connecting the measuring probes.
The characteristics of the reactor and the operating conditions for the experiments are
described in Table 3.A.1.
The overall set-up was monitored by a system of programmable relays (Bioline,
Inceltech, F-Toulouse).
Table 3.A.1. Operational data for the bioreactor
Parameter Value Material
Temperature 30 °C PT 100, Digital thermostat
pH 7 +/- 0.2 Mettler Toledo, CH-Greifensee
Dissolved Oxygen
(DO)
0.8-98 %
100 %=7.14 mg/l
Oximetric probe (L=16.4 cm)
Mettler Toledo, CH-Greifensee
Agitation 400 RPM 2 Rushton impalers
Aeration 5 l/min
Recirculation 13 l/min
KLa=34 h-1
Diaphragm pump, Haussman pumpen, CH-
Schlieren
65
Figure 3.A.1. Schematic presentation of the laboratory bioreactor
3.A.3.3. Analytical methods
Concentrations of CB and DCB were determined by capillary gas chromatography
(Varian Star 3400 Cx, Palo Alto, USA; Integrator C-R65A, Shimadzu, France), with a
flame ionization detector fitted and a DB-624 capillary column (30 m length, 0.53 mm
diameter) (J&W Scientific, Folsom, California). Samples were analyzed at a 5.8 ml/min
nitrogen flow rate and at 30 ml/min split injection. Injection and detector temperatures
were 220°C. Oven temperature was increased at a rate of 10°C/min from 80°C to a final
Water - jacket
NaOH 5M
CB, DCB
Motor
Drain or harv est
Oxygen probe (DO) pH electrode
Cultivation medium
Gas sample Liquid sample
i nlet air
Security valve (pressure <2 bars)
outlet air Condenser
Aeration
66
temperature of 145°C, which was maintained for 2.5 min. Retention times were 4.5 and
7.8 min for CB and DCB, respectively. The culture was sampled by means of a 1 ml Gas-
tight syringe (Dynatech, Serie A2, Bâton-Rouge, USA).
Chloride and ammonium ions concentrations were determined photometrically with a
flow injection analyzer (FIA, Tecator, Höganas, Sweden) at 463 and 590 nm,
respectively. Conductivity was determined directly in the culture medium with a
conductimeter (WTW, Weiheim, Germany). Concentrations of water-soluble metabolites
were measured by high performance liquid chromatography (HPLC) (Varian 910, Palo
Alto, USA), on a ORH801 column (Interaction chromatography Inc., San Jose, USA)
with sulfuric acid (0.01 N) as the mobile phase at a flow rate of 1ml min-1. Compounds
were detected by the IR refractivity index detector (Varian 9065 Polychrom, Palo Alto,
USA) and their retention times were compared with trans–trans muconic acid p.a.
quality, supplied by Fluka, CH-Buchs.
The accumulation of UV-absorbing metabolites was monitored by UV-spectrophotometer
(U-2000, Hitachi, Tokyo, Japan) at 255 nm after filtration (0.22 µm filter, Schleicher and
Schuell, G-Dassel). The dissolved organic carbon (DOC) concentration was measured
after filtration, HCl-acidification and CO2 elimination, using a TOC analyzer (Shimadzu,
Tokyo, Japan). The oxygen uptake rate was determined by measuring the concentration
of dissolved oxygen in the culture media using a polarographic probe (Ingold AG,
Urdorf, Switzerland) after reducing the agitation rate.
BOD5 – Biochemical Oxygen Demand – was measured during 5 days using an Hg free
WTW 2000 Oxitop© unit thermostated at 20 °C. Experiments were carried out according
to official protocol (Dir 92/69/EEC, O.J. L383A). The volume of the solutions poured in
the bottles was determined according to the theoretical oxygen demand without
nitrification (ThODNH3). It corresponded to 97 ml of medium.
The acute toxicity was assessed using a MicrotoxTM Model 500 Analyzer. This analyzer
is based on the use of a luminescent bacterium, viz. the strain Vibrio fischeri NRRL B-
11177. When properly grown, luminescent bacteria produce light as a by-product of their
cellular respiration. Any inhibition of cellular activity results in a decreased rate of
respiration and a corresponding decrease in the rate of luminescence. The toxicity is
67
determined by measuring, after 5 and 15 minutes of exposure, the concentration at which
50% of the light is lost because of the toxicity of the solution examined (EC50). The Toxic
Unit (TU) is used to express the data, as proposed by U.S. EPA (Lankford and
Eckenfelder, 1990): TU = 100/EC50.
The LIVE/DEAD® BacLightTM Bacterial Viability Kit L-7007 (Molecular Probes
Corporation, Canada) provides a quantitative index of bacterial viability. It uses a mixture
of SYTO® 9 green fluorescent nucleic acid stain which labels bacteria with both intact
and damaged membranes. Besides, the red fluorescent nucleic acid stain Propidium
Iodide penetrates only bacteria with damaged membranes. The bacterial samples were
centrifuged and filtered through 0.2 µm PC Memb 13 mm (Nucleopore CORP., Canada).
The Epi-Fluoro-Microscopic observation was carried out under UV light with emission
filters BP 450-490nm, reflector FT 510 and stop-filter LP 520 (Eclipse 800, Nikon,
Switzerland).
Acute toxicity immobilization tests were performed with Daphnia magna according to
OECD Guideline 202 (2004). Temperature was maintained at 20 °C and replicate groups
of 10 samples were carried out.
VOCs lost by stripping were continuously quantified by FID analyses during the process.
VOCs lost by stripping were taken into account for all the calculations.
The biomass concentration was monitored by dry-weight measurement according to
Dutch standard methods (NEN 32355.3) and by absorbance at 650 nm.
3.A.4. Results and discussion
3.A.4.1. Substrate Feeding Pattern
The process of substrate feeding by pulse (SPB) consists of injecting a small but precise
amount of CB/DCB mixture so that the global volume of the liquid medium remains
constant, while bacteria are allowed to mineralize the substrate.
SPB cultivation process is characterized by the successive sequences of pulses. The total
period of cultivation is determined by the overall behavior of the microbial community.
68
Usually bacteria are progressively inhibited by the accumulation of chlorides which
results from the degradation of the chlorinated substances or by the production of
intermediate compounds which are not converted into carbon dioxide and water.
The cultivation time consist of the repetition of a sequence which is called “pulse”.
Basically, one typical cycle started by aeration, until the medium was saturated by
oxygen (overall 5 minutes were needed to achieve the oxygen saturated conditions). Then
one pulse of substrate was added, with a special care to avoid CB and DCB stripping.
Then the substrate concentration in the medium sharply decreased until total substrate
disappearance.
A typical response of the process during one pulse is illustrated in Figure 3.A.2.
Figure 3.A.2. Disappearance of CB and DCB in the liquid phase, consumption of
Dissolved Oxygen, and evolution of absorbance at 255 nm during one pulse.
As shown in Figure 3.A.2, by introducing the compounds to the reactor at t = 0, the
oxygen concentration sharply decreased and remained constant even after disappearance
of CB/DCB. The consecutive low oxygen concentration was stated to be due to the
degradation of intermediate metabolites (León et al, 1999).
0
20
40
60
80
100
0 11 22 33
Time (min)
[CB
] et
[o
-DC
B]
(mg
/l), D
O (
%)
0
0.3
0.6
0.9
1.2
Ab
sorb
ance
250
nm
CB
o-DCB
DO
abs 255nm aeration
69
The maximum amount of substrate added by one pulse was determined by the available
oxygen concentration in the liquid medium. The pulse duration was the time between two
injections of substrate; it should be long enough to degrade the substrate and its
metabolites in order to avoid their accumulation in the liquid phase and consequently
their inhibitive effect.
During the first set of SPB, experiments were run with constant cycle time and loading
rates of 35 min and 125 mg l-1 of CB/DCB mixture, respectively. Then the optimal
biomass growth yield and productivity were achieved either by increasing the amounts of
substrates added by each pulse, or by reducing the cycle time. It should be noticed that
reducing cycle time results in a continuous approach.
In this study, both techniques (SPB and continuous feedings) were investigated and
compared. The results will be discussed below in more details.
3.A.4.2. Parameters of Control
The correct proceeding of the bacterial cultivation requires regular analyses of
concentration of metabolic intermediates, chlorides, ammonium, and cellular integrity in
the system.
3.A.4.2.1. Metabolic intermediates
UV absorbance at 255 nm (A255) measures the concentration of the aromatic
intermediates produced consequently to the degradation of the chlorinated solvents.
Seignez et al.12 presented A255 as a reliable parameter to control and optimize the
cultivation process. A255 can be indicative of the removal efficiency of the biosystems
and was measured regularly. Intermediates such as muconic acid, mono and
dichlorocatechols were analyzed by HPLC, using both columns, for aromatic and
aliphatic detection. Scheme 3.A.1 illustrates the most relevant pathways used to explain
CB and o-DCB degradation (Reineke and Knackmuss, 1984; Haigler et al, 1988).
Proteins and dissolved organic carbon were also quantified during the experiment. Table
3.A.2 presents 3 typical metabolic behaviors observed during batch and continuous
cultivation processes.
70
Table 3.A.2. Production of intermediate metabolites as the bacterial response to
continuous or batch cultivation processes. Substrate loadings are indicated in parenthesis
Cultivation
technique
Intermediate
productivity
(mgc l-1 h-1)
Intermediate
production
rate
(mgc gCB/dCB-1)
Maximum
intermediate
concentration
(mgc l-1)
Detected
metabolites
Toxicity
(DBO5,
Daphnia)
SPB
(200 mg l-1 h-1) 0.56 ± 0.05 10 ± 2 250 muconic acid No
SPB
(600 mg l-1 h-1) 7.5 ± 0.08 370 ± 20 600
3-chlorocatechol
dichlorocatechol No
Continuous
(200 mg l-1 h-1) 1 ± 0.05 20 ± 10 200
muconic acid
proteins (2 g l-1) No
As shown in Table 3.A.2, intermediate metabolites appear to be produced and accumulate
in all cases.
Even though stœchiometric chloride concentrations are attained, the complete
mineralization seems to be achieved lately by the further degradation of intermediates.
However, both productivity and production rates of intermediate compounds depend on
the substrate loading.
With the SPB process, increasing 3 times the amount of CB/DCB mixture injected by
pulse gives more than 30 times the production rates of metabolites.
When SPB process with 600 mg l-1 of CB/DCB was applied, metabolites tended to
accumulate in the system. After 8 hours (± 2) of total cycle run, 600 mgc l-1 of
intermediates was measured. This maximum intermediate concentration was concomitant
with the accumulation of CB/DCB in the system, the color of the cultivation medium was
observed to turn black and the acid addition rate needed for the pH controller sharply
increased. These observations were proved to characterize the typical inhibition of the
bacterial community (León et al, 1999).
71
Muconic acid was only detected in the 200 mg l-1 SPB and in the continuous processes,
but not with the 600 mg l-1 SPB technique.
It means that the metabolic pathway at 600 mg l-1 SPB process has already been stopped
before production of muconic acid. Indeed mono and dichlorocatechols accumulated (see
Scheme 3.A.1).
In this case, only 60% of stœchiometric chlorides were recovered.
Scheme 3.A.1. CB and o-DCB degradation routes achieved by Pseudomonas sp. (Haigler
et al, 1988)
The Daphnia bioassay showed that the intermediate compounds were not toxic. The
obtained cultivation media was even proved to be partially biodegradable, according to
O H
H
Cl O H
O H
Cl Cl
O H
O H
H
H
Cl O H
O H
H
H
Cl Cl
Cl
C O O H C O O H
Cl
C O O H C O O H
Cl
C O O H C O O H
O C O O H C O O H
O Cl
Cl Cl Cl
O 2 N A D H
N A D +
O 2
- C l -
c h l o r o b e n z e n e d i h y d r o d i o l s
c h l o r o c a t e c h o l s
muconic acids
c h l o r o b e n z e n e s
72
the DBO5 test. So industrial effluents could be pretreated with such a process and
consequently be rejected into a classical sewage treatment plant with no risk of
disruption.
Finally the two main factors which determine the viability of the overall cultivation
process are:
o the amount of substrate injected into the cultivation medium and
o the time needed for the bacteria to degrade intermediates.
If the two points are deficient, intermediate compounds tend to accumulate, which lowers
the removal of CB and DCB and provokes the inhibition of the CB/DCB conversion.
Consequently the bacterial community loses its ability to eliminate CB and DCB, and
cellular integrity is damaged.
Therefore the particular case of continuous feeding becomes relevant whenever the
amount of substrate approaches the optimal substrate loading determined with the SPB
process (i.e. 200 mg l-1 h-1).
In this context, high concentrations of proteins were detected (2 g l-1). No further
investigation was realized in this study but Part B in Chapter 3 will report the production
of Exo Polymeric Substances (EPS) by similar bacterial communities (Pseudomonas sp.)
which are able to degrade a mixture of toluene, ethylbenzene and xylenes.
These metabolites could protect the bacterial cells against the omnipresence of CB and
DCB. Indeed the two molecules as well as their metabolites of degradation are lipophilic
and toxic; they are known to be responsible of cellular and wall damages (De Bont,
1998).
3.A.4.2.2. Chloride concentration
The chloride ion concentration was determined either by FIA or conductivity of the
medium: both methods were satisfying.
Experimental measurements fitted well with the theoretical quantity of substrate added to
the medium and stœchiometric amounts of chlorides were released into the medium
during CB/DCB conversion.
73
This correct correlation confirmed the complete degradation of the parent molecules.
Since the measured values were compared with the FID analyses of the off-gases,
approximately 10 % of the injected substrate was estimated to be lost by stripping.
Oh and Bartha (1994) have reported that a chloride concentration higher than 7 g l-1 could
inhibit the bacterial growth. The above-mentioned threshold was systematically
overwhelmed in our experiments without disturbance. Indeed, during the SPB
experimentation tests, the chloride production rate was approximately 4 mg l-1 h-1 and the
maximal concentration was estimated to be 8.9 g l-1. Since higher substrates loadings
were supplied during the continuous cultivation process, the average production rate of
chlorides was at least 125 mg l-1 h-1 and the above-mentioned toxicity threshold was
passed after 95 h with a final concentration of chlorides of 11.8 g l-1.
3.A.4.2.3. Nitrogen contents
Ammonium concentration in the liquid phase was measured every day to avoid nitrogen
limitation for bacterial growth.
Nitrogen was supplied to the microorganisms by adding ammonium salts; nitrate
concentration in the reactor was 4.5 g l-1 and was never allowed to reach below 3 g l-1.
The overall consumption of ammonium was 150 mg gDW-1, and no significant difference
was observed between SPB and continuous techniques.
Nitrification was noticed once with the SPB process but not observed in the continuous
experiments.
3.A.4.2.4. Cellular Integrity
The cellular viability was investigated with the BaclightTM microscopic assay in the
course of cultivation process. This easy tool was particularly useful to optimize the cell
productivity.
Indeed, a high active cell ratio is indication of a high removal capacity. In this case it is
possible to increase the substrate loading rate, which permits to enhance the removal
capacity. On the contrary, when low active cell ratio is observed, substrate loading rate
should be decreased in order to avoid inhibition.
74
The overall evolution of the cellular viability was similar for both types of substrate
feeding (SPB or continuous).
A ratio of 70 % of healthy bacteria indicated that the process had reached a convenient
steady-state. In this context, even though the experiments were performed at constant
temperature and pH, bacteria showed to be resistant to unexpected short periods changes
of temperature (up to 40°C for about one day), or pH (4.5 during 10 hours).
3.A.4.3. CB and DCB Conversion
3.A.4.3.1. SPB Technique
Figure 3.A.2 is representative of a typical SPB degradation experiment during one pulse.
In the example, 10 minutes were necessary for the bacteria to degrade 100 and 25 mg l-1
of CB and DCB, respectively. CB disappeared immediately, whereas two steps were
observed before the complete removal of DCB.
In fact, different metabolic pathways are implicated to degrade CB and o-DCB (Haigler
et al, 1988; see Scheme 3.A.1).
Investigations on Pseudomonas sp. showed that growth on CB involves conversion of the
substrate to chlorocatechols via dioxygenase and diol dehydrogenase reactions.
Concerning o-DCB, it is converted by a dioxygenase to a hydrodiol, which is converted
to 3,4-dichlorocatechol. According to Sigel et al. (1992 ), a negative effect is exerted by
the two chlorides onto the aromatic ring, which slows down the overall reaction
compared to CB degradation.
During the first phase of substrate degradation, absorbance at 255 nm (A255) increased
from 0.6 to 0.9 due to the production of metabolic intermediates, and then decreased
progressively to the initial value when substrate and metabolites were totally degraded.
This was the criteria to start a pulse because A255 evolution could be linked with the risk
of inhibition due to toxic intermediate metabolites as well as initial molecules, and
absorbance is also easy to measure.
The oxygen uptake rate was measured by using the biological oxygen monitor (BOM) to
be 5.2 (± 0.5) g l-1h-1 at the start of the cycle during the VOCs uptake, due to the aerobic
degradation of the substrate.
75
Specific degradation rates of CB and DCB were investigated during the overall SPB
process.
The initial substrate uptake rate (qs,i) is calculated from the slope of the time course of the
CB/DCB removal from the gas phase (GC analyzes). Only the first 6 minutes after
injection of substrates were considered for this calculation.
The gas-liquid mass transfer coefficient was obtained from physical absorption of N2O as
described by Ebrahimi et al. (2003).
At 400 RPM, KLa was calculated to be about 0.015 s-1 and mass transfer rate was
calculated to be at least 4 times higher than the uptake rate of the compounds. So using
the concentration of the gas phase for initial uptake rate calculation is not a problem, but
by increasing biomass concentration and decreasing the compound concentration in the
gas phase, it is risky and mass transfer coefficient should be taken into account.
Figure 3.A.3 plots some typical slopes used to calculate the initial substrate uptake rates
for CB and o-DCB. Global substrate uptake rate (qs,g) was calculated based on the total
amount of CB/DCB removed during one pulse (by correcting for stripping). As qs,g also
took into account the time needed for degradation of both CB/DCB and intermediate
metabolites, it was lower than the initial rate (qs,i).
Figure 3.A.3 shows the typical time course of the substrate uptake rates during the SPB
cultivation process.
Initial substrate uptake rates were maximal at the beginning of the cultivation process and
sloped down to a mean value of 0.07 and 0.02 g gDW-1 h-1 for CB and DCB, respectively.
The same behavior has already been reported (Bielefeldt and Stensel, 1999a; Bielefeldt
and Stensel, 1999b).
In the case of chlorinated solvents, the substrate removal can be directly correlated with
the chloride released in the cultivation medium during the reaction phase. Note that this
should be distinguished from the full mineralization of the substrate.
In the present study, stœchiometric amounts of chlorides were always completed, except
before an observed inhibition of the bacterial community, due to an excess of substrate
loading. Thus, as the release of chlorides happens during the last steps of CB/DCB
76
degradation (see Scheme 3.A.1), it means that substrate was degraded and not just
absorbed inside the bacterial cell.
Therefore decreasing qs reflects a metabolic inhibition and not an increasing difficulty for
the bacteria to incorporate the substrate.
Figure 3.A.3. Calculation and evolution of the degradation rates qs during the cultivation
process
3.A.4.3.2. Continuous Feeding
System was operated in a fed-batch mode with a continuous substrate feeding.
Change in the total liquid volume was negligible because of dosing a concentrated
substrate mixture.
During the continuous cultivation process, the CB/DCB loading was progressively
increased along the time course and ranged between 200 and 520 mg l-1 h-1.
Consequently to every load changes, at least one hour was necessary for the biological
system to reach a steady-state. For instance, with 430 mg l-1 h-1 of CB/DCB injected
(t = 3 d)
y = -20.32x + 98.93
R2 = 1.00
y = -11.14x + 100.49
R2 = 0.99
0
20
40
60
80
100
0 2 4 6 8 10
Time (min)
Con
cent
ratio
n (%
)
CB
DCB
(t = 5 d)
y = -17.59x + 99.77
R2 = 0.99
y = -9.59x + 102.51
R2 = 0.99
0
20
40
60
80
100
0 2 4 6 8 10
Time (min)C
once
ntra
tion
(%)
(t = 7 d)
y = -8.89x + 100.06
R2 = 0.99
y = -6.05x + 99.86
R2 = 0.99
0
20
40
60
80
100
0 5 10Time (min)
Con
cent
ratio
n (%
)
(t = 11 d)
y = -13.07x + 101.17
R2 = 1.00
y = -7.24x + 99.23
R2 = 0.99
0
20
40
60
80
100
0 2 4 6 8 10Time (min)
Con
cent
ratio
n (%
)
77
inside the bioreactor, the measured residual substrate concentration stabilized to 3.6 and
7.6 mg l-1 of CB and DCB, respectively.
Along with the microbial adaptation to the CB/DCB degradation, the residual
concentration was systematically observed to decrease until zero.
However 520 mg l-1 h-1 of CB/DCB was found to be the threshold for the community to
simultaneously degrade CB and DCB. Indeed, above this load (corresponding to 100 mg
l-1 h-1 of DCB), 7.4 mg l-1 of DCB remained into the liquid phase; this residual
concentration being stable. Higher loads provoked the inhibition of the both substrate
degradation.
As well as for SPB process, substrate uptake rates decreased in the course of continuous
cultivations. Obtained maximal global CB and DCB uptake rates reached respectively 53
and 13 mg gDW-1 h-1 for continuous feeding, versus 61 and 14 mg gDW
-1 h-1 for SPB
technique.
The above-mentioned data were the maximum substrate uptake rates which could be
obtained, regardless of the cultivation period. Regarding these values, the two cultivation
processes appeared to be comparable on kinetics aspects.
3.A.4.4. Biomass Yield and Productivity
Biomass yield and productivity were determined and compared for both SPB and
continuous feeding techniques (Table 3.A.3). SPB was performed with either pure
oxygen or air during the aeration step (See section 3.A.4.1).
Every parameter presented in Table 3.A.3 shows that the continuous feeding of substrate
was the most productive approach for bacterial cultivation: the obtained biomass
productivity (100 mgDW l-1 h-1) was twice the global productivity conducted with the SPB
technique.
Both SPB processes with air and pure oxygen were found to produce similar biomass
contents regarding time (global productivity averaged 50 mgDW l-1 h-1), but not regarding
the substrate loading. Thus considering global yields, oxygenated SPB was comparable to
the continuous process (i.e. 400 mgDW gCB/DCB-1).
78
Table 3.A.3. Biomass yields and productivities achieved by cultivation processes
operated within aerated SPB, oxygenated SPB or continuous feeding techniques.
SPB (air) SPB (O2) Continuous
Maximal productivity
(mgDW l-1 h-1) 66 ± 6 56 208 ± 6
Global productivity
(mgDW l-1 h-1) 48 ± 2 51 100 ± 3
Global yield
(mgDW gCB/DCB-1)
280 ± 8 400 ± 25 360 ± 20
Maximal substrate loading
(mgCB/DCB l-1 h-1) 240 155 520
Optimal substrate removal rate
(mgCB/DCB l-1 h-1) 235 ± 5 100 ± 20 470 ± 6
Maximal biomass concentration
(gDW l-1) 13 8 20
However, the continuous feeding mode could perform at least twice the maximal
substrate loading that SPB could carry out. Moreover the continuous system was more
relevant compared to the others because its optimal substrate removal rate was twice the
optimal rate obtained with the SPB process operated with air and five times the optimal
substrate removal rate of SPB operated with oxygen.
Hence biomass production of the continuous reactor was higher than for sequential batch
reactors.
This behavior was surprising, due to the toxicity of CB and o-DCB.
In particular, MicrotoxTM acute toxicity testing was performed on these substances. The
obtained EC50 (5 min) were 16.94 (± 0.9) and 3.13 (± 0.2) mg l-1 for CB and o-DCB,
respectively, which is in accordance with literature (Bazin, 1985; Blum and Speece,
1991; Warne et al., 1999).
79
Besides, these substances and some of their metabolic byproducts are known to damage
the bacterial membranes and provoke inhibition (De Bont, 1998).
This is the main reason why SPB (Substrate Pulse Batch) processes were aimed for
treating VOCs, such as BTEX and halo-aromatics (León et al, 1999; Shim and Yang,
1999).
Therefore it was risky to test the continuous feeding of CB and o-DCB.
First, it allowed residual amounts of CB and o-DCB, and their possible accumulation
nearby and inside the cells.
Second, it favored the use of the parent molecules instead of the subsequent intermediate
metabolites. In this case, it was shown that the metabolic degradation chain could be
blocked (Klecka and Gibson, 1981).
In the present study, these phenomena were avoided. It was certainly due to the correct
adaptation of the bacterial community able to degrade CB and o-DCB, and also because
the substrate loadings were progressively increased in the course of the cultivation
process.
3.A.5. Conclusion
Active bacterial community was produced to degrade chlorinated solvents. Four regular
analyses were used as reliable parameters in order to control bacterial growth and
CB/DCB degradation. They consist of UV absorbance at 255 nm (intermediate contents),
chloride production, nitrogen consumption and cellular viability.
Biomass yield and productivity as well as the CB/DCB degradation rates were
determined and compared for both SPB and continuous techniques. Between the two
experimented processes, the continuous feeding of substrate was found to enhance both
biomass productivity and CB/DCB removal capacity.
However, two different endpoints can be considered. First, a biotechnological approach
tends to maximize both biomass productivity and global yields. In this case, the
continuous substrate feeding process was the most relevant, as the obtained values were
80
more than twice those obtained for the SPB process (360 mg of dry biomass produced per
g of CB/DCB consumed versus 280 mgDW gCB/DCB-1). The second strategy concerns either
environmental or industrial topics and focuses the maximal removal capacity of the
biotreatment. Substrate degradation rates are to be optimized so that the process allows a
sustainable treatment of polluted gaseous effluents (i.e. efficient and stable). In this case,
the continuous feeding process was also the most relevant technique with 470 ± 6 mg
CB/DCB l-1 h-1. Finally, the continuous supply of substrate corresponds to an enticing mode
of biomass production regarding industrial prospects and it is usually easier to operate
compared to batch processes.
3.A.6. Nomenclature
A255 = abs 255 nm = absorbance at 255 nm
BOD = biochemical oxygen demand
DO = dissolved oxygen
DW = dry weight
KLa = mass transfer coefficient
qs,g = global substrate uptake rate
qs,i = initial substrate uptake rate
RPM = rotation per minute
SPB = substrate pulsed batch
TEX = toluene, ethylbenzene, o-, m-, p-xylene
3.A.7. References
Pearson, C.R., 1982. The Handbook of Environmental Chemistry, Anthropogenic
Compounds; Vol. 3, Part B; Springer: New York, pp. 89.
Oh, Y.S., Bartha, R., 1994. Design and performance of a trickling air biofilter for
chlorobenzene and o-fichlorobenzenze vapors. Appl. Environ. Microbiol. 60, 2717-
2722.
81
Barona, A., Elías, A., Arias, R., Cano, I., Gonzales, R., 2004. Effect of sudden variations
in operating conditions on biofilter performance. Biochem. Engin. J. 1, 25-31.
Kennes, C., Thalasso, F., 1998. Waste gas biotreatment technology. J. Chem. Technol.
Biotechnol. 72, 303-319.
Kennes C., Veiga, M.C., 2001. Conventional biofilters. In Bioreactors for Waste Gas
Treatment; Kluwer Academic Publishers: Dordrecht, The Netherlands, pp. 47-98.
Lankford, P., Eckenfelder Jr., W., 1990. Toxicity Reduction in Industrial Effluents. Van
Vostrand Reinhold, New York, USA.
Prado, O.J., Eiroa, M., Veiga, C., Kennes, C., 2003. Bioreactors for the treatment of
industrial waste gases containing formaldehyde and other aliphatic compounds. In
Biotechnology for the Environment: Wastewater Treatment and Modelling, Waste Gas
Handling, Kluwer Academic Publishers: Dordrecht, The Netherlands, pp. 259-273.
Prado, O.J., Veiga, M., Kennes, C., 2005. Treatment of gas-phase methanol in
conventional biofilters packed with lava rock. Water Res. 39, 2385-2393.
León, E., Seignez, C., Adler, N., Péringer, P., 1999. Growth inhibition of biomass
adapted to the degradation of toluene and xylenes in mixture in a batch reactor with
substrates supplied by pulses. Biodegradation 10, 245-250.
Seignez, C., Vuillemin, A., Adler, N., Péringer, P., 2001. A procedure for production of
adapted bacteria to degrade chlorinated aromatics. J. Hazard. Mater. B84, 265-277.
OECD Guideline 202, 2004. Guidelines for Testing of Chemicals: Daphnia magna
Reproduction Test.
Seignez, C., Atti, A., Adler, N., Péringer, P., 2002. Effect of biotrickling filter operating
parameters on chlorobenzenes degradation. J. Environ. Eng. 128, 360-366.
Reineke, W., Knackmuss, H.J., 1984. Microbial Metabolism of Haloaromatics: Isolation
and Properties of a Chlorbenzenze-Degrading Bacterium. App. Environ. Microbiol. 47,
395-402.
Haigler, B., Nishino, S., Spain, J., 1988. Degradation of 1,2-Dichlorobenzene by a
Pseudomonas sp. Appl. Environ. Microbiol. 54, 294-301.
82
De Bont, J., 1998. Solvent-tolerant bacteria in biocatalysis. Biotech. 16, 493-499.
Sigel, H., 1992. Metal Ions in Biological Systems: Degradation of Environmental
Pollutants by Microorganisms and their Metalloenzymes. Metal Ions in Biological
Systems Vol. 28, Marcel Dekker: New-York, pp. 157-192.
Ebrahimi, S., Kleerebezem, R., Van Loosdrecht, M., Heijnen, J., 2003. Kinetics of the
reactive absorption of hydrogen sulfide into aqueous ferric sulfate solutions Chem. Eng.
Sci. 58 (2), 417-427.
Bielefeldt, A., Stensel H., 1999a. Evaluation of biodegradation kinetic testing methods
and longterm variability in biokinetics for BTEX metabolism. Wat. Res. 33, 733-740.
Bielefeldt, A., Stensel H., 1999b. Modeling competitive effects during biodegradation of
BTEX mixtures. Wat. Res. 33, 707-714.
Bazin, C., 1985. Sensibilité comparée des tests Daphnie et Microtox pour 14 substances à
risque toxique élevé. DEA, Université Claude Bernard, Lyon I.
Blum, D., Speece, R., 1991. Quantitative structure-activity relationships for chemical
toxicity to environmental bacteria. Ecotoxicol. Environ. Saf. 22 (2), 198-224.
Warne, M., Osborn, D., Lindon, J., Nicholson, J., 1999. Quantitative structure-toxicity
relationships for halogenated substituted-benzenes to Vibrio fischeri, using atom-based
semi-empirical molecular-orbital descriptors. Chemosphere 38 (14), 3358-3382.
Shim, H., Yang, S., 1999. Biodegradation of benzene, toluene, ethylbenzene, and o-
xylene by a coculture of Pseudomonas putida and Pseudomonas fluorescens
immobilized in a fibrous-bed bioreactor. J. Biotechnol. 67 (2-3), 99-112.
Klecka, G., Gibson, T., 1981. Inhibition of 2,3-dioxygenase from Pseudomonas putida by
3-chlrocatechol. Appl Environ. Microbiol., 41, 1159-1165.
83
CHAPTER 3.
Biodegradation of Volatile Organic Compounds
PART B.
MASS PRODUCTION OF BACTERIAL COMMUNITIES ADAPTED TO THE DEGRADATION OF VOLATILE
ORGANIC COMPOUNDS (TEX)
3.B.1. Abstract
This study focuses on the mass cultivation of bacteria adapted to the degradation of a
mixture composed of toluene, ethylbenzene, o-, m- and p-xylenes (TEX). For the
cultivation process Substrate Pulse Batch (SPB) technique was adapted under well
automated conditions. The key parameters to be monitored were handled by LabVIEW
software including, temperature, pH, dissolved oxygen and turbidity. Other parameters,
such as biomass, ammonium or residual substrate concentrations needed offline
measurements.
SPB technique has been successfully tested experimentally on TEX. The dynamic
behavior of the mixed bacterial population was observed and discussed under different
operational conditions. Carbon and nitrogen limitations were shown to affect the integrity
of the bacterial cells as well as their production of exopolymeric substances (EPS).
Average productivity and yield values reached 0.45 kgDW m-3 d-1 and 0.59 gDW gC-1,
respectively. Obtained data come up to the industrial specifications and indicate the
feasibility of controlled SPB technique.
Submitted to Biodegradation
84
3.B.2. Introduction
Toluene, ethylbenzene, and xylene (TEX) are among the most important contaminants
present in surface and groundwater which usually originate from the leakage of tanks
containing petroleum-derived products and industrial wastewaters (Akunar-Askar et al.,
2003).
Because they are both toxic and relatively water soluble compared with other petroleum
constituents, their entry into surface and drinking water is of major concern. In spite of
governmental intervention in many countries, their emission to the environment is still
escalating.
Therefore, there is a need to develop and optimize technologies for removing TEX from
groundwater, especially when downstream drinking water resources are a concern.
Among all remediation technologies for treating TEX-contaminated groundwater,
bioremediation appears to be an economical, energy-efficient and environmentally sound
approach.
Microorganisms are able to degrade TEX under aerobic, microaerobic or hypoxic, as well
as anaerobic conditions (Villatoro-Monzon et al. 2003). Development and upgrading of
the biotechniques for simultaneous and efficient removal of all TEX with a mixed culture
are of major challenges.
The substrate pulse batch (SPB) technique is attractive for the production of a specialized
biomass used to inoculate biosystems which treat volatile organic compounds (VOCs).
SPB technique is a batch cultivation process in which substrates are supplied in pulses.
This technique can be considered as a special limit case of a conventional, semi-
continuous extended culture operation: total volume remains virtually constant because
substrate is fed in a gaseous, solid or very concentrated liquid form, according to an
intermittent profile.
This technique aims to save both heterogeneity and kinetic dynamics of the mixed
population thanks to the continual variation of substrate concentration in the medium.
Short contacts time due to rapid degradation also avoid toxic damages which may result
from the solvent effects onto bacterial cells (De Bont 1998). This process is
85
recommended for achieving a more stable and efficient biomass culture (León et al. 1999;
Seignez et al. 2001). However, industrial implementation of SPB technique is limited
because of its rules of thumb.
This study was aimed to improve the productivity and yield of a particular microbial
culture for degrading a TEX mixture using SPB technique. This research was in context
of an European Project, “Bioreactor For Innovative Mass Bacteria Culture, BIOMAC”
(see: www.eureka.be, project E!2497). This study presents the kinetics of an adapted
bacterial consortium to degrade TEX in an automated bioreactor using SBP technique.
3.B.3. Materials and methods
3.B.3.1. Inoculum Preparation
The bacterial consortium used to inoculate the bioreactor was selected in our laboratory
in a previous work (León et al. 1999). The bacteria were originally issued from the sludge
of a wastewater treatment plant (Novartis and Rohner AG, Basel, Switzerland) and from
the biotrickling filters (Rohner AG, Basel, Switzerland), adapted to the TEX degradation
and cultivated on TEX or toluene as sole source of carbon then preserved at -80°C.
A sample of this adapted mixed culture was cultivated in flasks containing nutrient
medium (CM67, Oxoid LTD, Basingtoke, England) at 30°C for 1 day.
Then the obtained biomass was collected by centrifugation and transferred into the batch
reactor.
The nutrient medium consisted of inorganic salts and vitamins necessary for the growth
of microorganisms, the same composition as previously reported in Chapter 3, Part A,
with ammonium as the nitrogen source.
In all the experiments, the sole carbon and energy source consisted of 90 % (v/v) toluene
(Tol), 2.5 % ethylbenzene (EB), 6.5 % m- and p-xylenes (X), 1.3 % o-xylene. This
mixture is called “TEX”. Compounds were supplied by Fluka, Buchs, Switzerland.
Experiments were made under non-sterile conditions. TEX were supplied by a precision
feeding pump (SpectraPhysics, San José, USA).
86
3.B.3.2. Bioreactor and Automation Program
The experimental set-up used in this study is presented in Figure 3.B.1.
Bioreactor is made of stainless steel and Pyrex glass; with a total volume of 14.5 l. The
working liquid volume was 10 l. Numerous ports along its height and at the bottom allow
feeding, sampling and connecting the measuring probes.
The characteristics of the reactor and the operating conditions for the experiments are
described in Table 3.B.1.
Table 3.B.1. Operational data for the bioreactor
Parameter Value Material
Temperature 35 °C PT 100, Digital thermostat
pH 6.5-7 Electrode, Liquisys S, Threshold contact e
Dissolved Oxygen
(DO)
0.8-98 %
100 %=7.14 mg/l
Oximetric probe (L=31.7 cm)
Ingold Messtecknick
Agitation 170-800 RPM Motor : Lust
Aeration 0.5-0.7 l/min
KLa=3-9 h-1 Brooks Instrument
The overall set-up was controlled by the so-called BioOPT program. This LabVIEW
computer-based monitoring program allows data acquisition and was used to optimize the
biodegradation process.
Modulable FieldPoints (National Instruments) were the connecting tools between
computer and instrumental systems. The main electrical components are listed in Table
3.B.2.
Table 3.B.2. List of the National Instruments modules
FP-2000 Ethernet network module capable of running LabVIEW Real-Time
embedded code and 3 Mb Flash Memory for user
FP-AI-110 8 analog inputs 0-20 /4-20 mA, 16 bits, 50/60 Hz filtering
FP-AO-200 8 analog outputs 0-20 /4-20 mA, 12 bits
FP-RLY-420 8 relays, SPST (form A), 3 A at 250 VAC or 35 VDC
FP-RTD-PT100 Dual Channel I/O: 4-wire RTD input modules (-50 to 350 °C)
FP-TB-10 Dual-Channel terminal base for 6 dual-channel I/O modules
87
Figure 3.B.1. Schematic presentation of the laboratory bioreactor
3.B.3.3. Analytical methods
Concentrations of TEX mixture were determined by capillary gas chromatography (GC-
14B, Integrator C-R65A, Shimadzu, France), with a flame ionization detector fitted and a
DB-624 capillary column (30 m length, 0.53 mm diameter) (J&W Scientific, Folsom,
California). Samples were analyzed at a 32 ml/min nitrogen flow rate and at 30 ml/min
split injection. Injection and detector temperatures were 220 and 300°C, respectively.
Oven temperature was increased at a rate of 4°C/min from 85°C to a final temperature of
Water-jacket
NaOH 5M
TEX
Motor
DO-2 probe
Drain or harvest
Turbidity probe
pH electrode
Cultivation medium
DO-1 probe
Gas sample
Liquid sample
inlet air
Security valve (pressure <2 bars)
outlet air Condenser
88
110°C. The culture was sampled by means of a 250 µl Gas-tight syringe (Dynatech, Serie
A2, Bâton-Rouge, USA).
Volatile Organic Compounds (VOCs) lost by stripping were continuously quantified by
FID analyses during the process. The amounts were systematically taken into account for
any calculation of the TEX conversion rate.
Dissolved organic carbon concentration was measured by an IR detector Shimadzu TOC-
500 (Burkard Instrumente, Switzerland). The reactor sample was filtrated at 0.45 µm
(Schleicher&Schuell, Germany), acidified by HCl 2 M and the CO2 was purged before
analysis.
Metabolites estimation was made by measuring the optical density of the filtered sample
at 255 nm in an UV spectrophotometer (Hitachi U-2000, Tokyo, Japan). Ammonium and
nitrate in the liquid phase of the batch reactor were determined by the enzymatic method
(Boehringer, Manheim, Germany) to control the nitrogen measurement for TEX
degradation.
The biomass concentration was monitored by dry-weight measurement according to
Dutch standard methods (NEN 32355.3) and by absorbance at 650 nm.
The LIVE/DEAD® BacLightTM Bacterial Viability Kit L-7007 (Molecular Probes
Corporation, Canada) provides a quantitative index of bacterial viability. It uses a mixture
of SYTO® 9 green fluorescent nucleic acid stain which labels bacteria with both intact
and damaged membranes. Besides, the red fluorescent nucleic acid stain Propidium
Iodide penetrates only bacteria with damaged membranes. The bacterial samples were
centrifuged and filtered through 0.2 µm PC Memb 13 mm (Nucleopore CORP., Canada).
The Epi-FluoroMicroscopic observation was carried out under UV light with emission
filters BP 450-490nm, reflector FT 510 and stop-filter LP 520 (Eclipse 800, Nikon,
Switzerland).
Extracellular polymeric substances (EPS) were extracted as follows. Samples of the
culture (10 ml) were centrifuged at 16.000 x g for 30 min at 4°C. The supernatant
polysaccharides (dissolved EPS) were isolated by precipitation with 3 volumes of
ethanol, and kept overnight at 4°C (Duenas et al. 2003). The pellets (bound EPS +
89
bacteria) were resuspended in nutrient medium containing 10 mM EDTA (Platt et al.
1985). After sedimentation of the bacterial cells by centrifugation, the supernatant was
treated as for the dissolved EPS. After centrifugation of both bound- and dissolved-EPS
tubes (5.000 x g, 20 min, 4°C), the pellets were dispersed in aqueous 80 % ethanol and
centrifuged again (3 times). The final precipitates were dissolved in distilled water (50
ml). The carbohydrate content was determined by the Anthrone method modified from
Standard Methods (Raunkjauer et al. 1994) with glucose as the standard.
The determinations were done in triplicate and the average values were calculated with
less than 5 % standard deviation.
3.B.4. Results and discussion
3.B.4.1. Bioengineering
To automate the production of adapted bacteria able to degrade TEX, a LabVIEW
computer-based monitoring program was developed for command and data acquisition.
First, the key parameters implicated in the biodegradation process are presented. Then
schedules are proposed to reinforce the interaction between command and control in
order to provide a completely autonomous system.
3.B.4.1.1. SPB technique for biomass cultivation
SPB technique has already been explained and illustrated in Chapter 3, Part A for
chlorobenzene and dichlorobenzene biodegradation. In Part B, the target VOCs are TEX.
SPB technique aims to save both overall heterogeneity and kinetic dynamics of the mixed
population thanks to the continual variation of substrate concentration in the medium.
Short contacts due to rapid degradation also avoid toxic damages which may result from
the solvent effects onto bacterial cells (De Bont, 1998). Therefore such a process allows a
better stability to the culture and improved biomass productivity.
90
However it is mainly built on empiricism. In this context, the operator plays a crucial
role, which limits the industrial implementation. This is the reason why this study focuses
on the automation of the process.
In this study, the cultivation of mixed bacteria was handled in a batch culture with the
pulse feeding technique. The overall process can be categorized into two periods.
o First, microorganisms are progressively fed with increasing but still low amounts
of TEX. This step allows the induction of the enzymatic pool and strongly influences the
further TEX degradation (Grady et al., 1996).
o Second, the most efficient substrate loading is investigated in order to improve
both productivity and degradation yields.
During the cultivation, four parameters are checked to have the best conditions for
microbial growth:
o Concentrations in the gas phase of TEX compounds;
o Concentration in the liquid phase of the produced intermediate compounds;
o Oxygen concentration in the liquid phase. It should be at least 20 % of the
saturation concentration. It is measured in two places in the bioreactor (Dissolved
Oxygen probes DO-1 and DO-2);
o Nitrogen concentration (ammonia) in the batch medium. It shall be higher than
100 mg/l to avoid limitation of the bacterial growth.
In this study, volatile substrates (TEX) are injected by repeated pulses in the gaseous
phase. 0.15 m above the liquid surface is the selected height to perform that crucial step:
it favors and regulates the injection of the convenient amount of solvents (no
overpressure) and limits any uncontrolled volatilization losses (stripping process).
The equilibrium between gas and liquid phases is reached in 2 minutes after the end of
the TEX injection. The stripping process depends on both aeration and agitation.
91
Three steps are successively performed:
o Phase 1 – aeration, for about 5 minutes to saturate the liquid phase.
Air flow was set at 7 to 10 l/min. If TEX were injected at the beginning of this phase,
80% of the substrate would be stripped.
o Phase 2 – injection of TEX, within 15 seconds.
The substrate amounts provided to the culture can be modified according to the
magnitude and/or the frequency of the injected pulses.
o Phase 3 – reaction.
During this phase, biodegradation of the parent molecules and their intermediate
metabolites occur. The time period allowed for this phase depends on the microbial
activity.
During phases 2 and 3, the air flow was set at 0.5 l/min to minimize the stripping
(evaluated at respectively 2% and 7 to 23% of the initial TEX amount).
3.B.4.1.2. Automation parameters
Each of the three above-mentioned phases can be precisely defined at the beginning of
the cultivation process using a basic LabVIEW computer-based monitoring program.
Experimental data are thus automatically recorded, so that the instrumental system is
completely autonomous, except for the decision to increase the substrate loading.
Indeed, the increase in substrate loading depends on the biomass behavior. It is difficult
to forecast the substrate needed, because it would mean that bacteria should consume the
solvents as expected.
In fact, bacterial consumption means (1) incorporation of the substrate, and only after (2)
degradation of the parent molecules and the intermediate metabolites. In particular,
aerobic biodegradation of TEX is correlated with oxygen consumption.
Experimentally, a minimum concentration of dissolved oxygen (DOmin) is recorded when
no residual concentration of TEX is measured in the gaseous phase (manual GC
analyses).
92
Therefore it is possible to indirectly control the accurate degradation of TEX by the
automatic and continuously recorded DO evolution.
Hence, two parameters can be used to characterize the effectiveness of the bacterial
cultivation:
o time necessary to degrade TEX during phase 3,
o decrease of DO.
First, time delay of phase 3 can be forecasted according to the cellular yield of biomass
production YX/S. Indeed, the global rate of substrate degradation qs (qs= µ/YX/S) may be
calculated. Then the frequency of pulses is determined so that the charge of TEX is
inferior to the maximum degradation capacity of consortium (= qs.[X]). Unfortunately,
this method is limited by the evolution of the consortium characteristics during the
cultivation. Moreover, concentration of biomass still needs to be determined (manual
operation).
Second, as previously mentioned, concomitant consumption of oxygen and TEX permits
to regulate the overall process according to the variations of DO values. Indeed when
DOmin is reached, it means that no more substrate is available for bacteria. Hence, a new
pulse of substrate can be injected.
However an extra-delay of time appears necessary after having reached DOmin. This
period of time corresponds approximately to 1/3 of the delay necessary for the TEX to
disappear from the aqueous phase; i.e. 1/3 of the time necessary to reach DOmin.
This delay is necessary for the degradation of the intermediate by-products inside the
bacterial cells.
By this way, the succession of pulses can be automatically monitored and adapted to the
real progression of the cultivation.
Monitoring of the cultivation process by the DO variations implies the reliability of DO
data. However, DO measurements can be faltered in two circumstances.
The first case happens when biomass growth implies an increasing O2 demand: O2
transfer becomes limiting into the liquid phase if the agitation remains the same. Thus
DO variation is attenuated and no DOmin is clearly noted (Figure 3.B.2a).
93
It outlines the need to link DO variation and agitation. For eg., ∆DO > 50 % can be
specified in BioOPT program. Underneath this value, agitation would be automatically
increased.
The second case happens when microorganisms become unhealthy: O2 consumption
decreases but shortening the cycles would favor the bacterial inhibition. On the contrary,
phase 3 should be lengthened (Figure 3.B.2b).
In both cases, it is supposed that DO probes are cleaned regularly (no bacterial film).
Figure 3.B.2. Evolution of DO during automatic runs of biomass cultivation
The obtained results indicate that using SPB method under automated conditions
(computer control) gives at least similar results to those obtained with an operator’s
supervision (unpublished results).
3.B.4.2. Microbiology
To optimize the production of adapted bacteria able to degrade TEX according to
industrial specifications, the cultivation of a bacterial community was carried out under
different operational conditions and kinetics features were followed.
0
10
20
30
40
50
60
70
0 1 2 3 4 5
Time (h)
DO
(%)
DO2 [%]DO1 [%]
limiting agitation
increased agitation
0
20
40
60
80
100
120
0 3 6 9 12 15 18Time (h)
DO
(%)
progressive inhibition of the
consortium
(a) (b)
94
3.B.4.2.1. Microbial growth
Several cultivations were carried out in order to automate the process. The key point was
to meet a convenient relation between (i) the amount of solvents which are the carbon
source and (ii) their toxic effects, which inhibit the bacterial growth.
For cultivation process, SPB was adopted because this technique allows the bacteria to
degrade a low amount of the initial molecule into less toxic metabolites. Each pulse of
substrate favors the bacterial growth and the acclimation to the solvents.
Higher substrate amounts are progressively introduced into the biological system. It was
shown that low amounts of TEX (326 mg l-1 every 2 h during 6 days) were necessary to
acclimate bacteria to TEX. After this step, the shorter the lag phase, the more resistant the
bacterial consortium to high substrate loading (696 mg l-1 every 5 minutes for instance).
Evolution of biomass concentration and substrate mass conversion rate during the overall
production process were investigated. Figure 3.B.3 illustrates the time course of the
increase on biomass concentration: 23 g/l of biomass was produced within 17 days.
Figure 3.B.3. Evolution of the biomass concentration during the cultivation process
During this period approximately 0.8 l of TEX mixture was injected into the biological
system. According to the carbon contents of the off-gases which were continuously
0
5
10
15
20
25
0 3 6 9 12 15 18
Time (d)
Bio
mas
s co
ncen
trat
ion
(g/l)
95
measured (FID analyses), 6.5 % of the loaded TEX was stripped. Total biomass yield and
biomass productivity were estimated to be 0.3 gDW gTEX-1 and 0.06 gDW l
-1 h-1, respectively.
Specific growth rates (µmax) were calculated to be about 0.007 h-1. Overall, calculated
growth rates ranged between 0.003 and 0.011 h-1 depending on carbon or nitrogen
contents in the nutrient medium. These values are comparable with other reports. For
instance Shim and Yang (1999) have reported that specific growth rates ranged between
0.001 and 0.002 h-1 with only toluene or o-xylene as substrate.
After the initial period of induction (6 days) and during 30 days of cultivation, interval of
time between successive pulses was reduced from 1.5 h until 0.5 h.
Besides and alternatively to the higher frequency of pulses, the amounts of TEX were
also increased from an initial concentration in the liquid phase of 24 mgc/l until a final
concentration of 67 mgc/l (stripping losses already subtracted). Therefore average
productivity and yield values reached 0.45 kgDW m-3 d-1 and 0.59 gDW gc-1, respectively.
These data come up to the industrial specifications and confirm the benefits of such an
improved technology.
3.B.4.2.2. Impact of C and N concentration on cell viability
The cellular viability was investigated during the cultivation process. After inoculation,
the ratio of healthy cells was approximately 50 %. It reached a climax (90 %) after 24
hours and then progressively decreased to 60 % within 16 days. Besides, it was shown
that limiting the amounts of carbon and nitrogen provoked cellular damages and was
concomitant with a lower productivity. This effect was reversible, since higher N and C
loadings allowed the cells to recover their integrity: this bacterial property is called
resilience.
Exo-Polymeric Substances (EPS) were also studied at different cultivation conditions.
Table 3.B.3 shows that limiting C or N contents of the nutrient medium change both the
composition and the total amount of EPS.
For instance, when the C source was limiting, the total amount of EPS was twice the
amount measured with the adequate supply of C and N.
96
Besides, limiting N provoked an increase of carbohydrates contents (92 % instead of 40
% with balanced C and N contents in the nutrient medium).
Since EPS are predominantly composed of carbohydrate and protein, the ratio between
each constituent depends on the relative C and N amounts in the nutrient medium.
Their active secretion is also linked to the environmental conditions: carbon starvation
provokes the production of dissolved EPS as a response to an energetic deficiency. On
the contrary, a nitrogen limited culture tends to minimize the overall EPS production and
certainly favors its consumption.
The similar link between nutrients, bacterial viability and growth has been demonstrated
by Durnaz and Sanin (2001) and Liu and Fang (2003).
Table 3.B.3. Influence of C and N on the cellular metabolism of the mixed population.
Conditions Limiting C Limiting N Balanced C and N
Viability 30 % 50 % 65 %
Dissolved EPS Carbohydrates contents
2.5 g/gDW
40 %
0.2 g/gDW
92 %
1.0 g/gDW
40 %
Bound EPS Carbohydrates contents
0.8 g/gDW
25 %
0.1 g/gDW
67 %
0.5 g/gDW
40 %
3.B.4.2.3. Substrate conversion
Typical kinetics of degradation of TEX during one pulse is presented in Figure3.B.4.
Three steps can be distinguished during elimination of TEX after injection into the liquid
phase:
o fast removal of TEX, followed by an increase of oxygen consumption rate during
the first 15 minutes,
o slower removal of TEX and stabilization of DO (lower bacterial consumption),
o no residual concentration of TEX, significant increase of both DO and A255
(absorbance at 255 nm). This step should correspond to the exportation of intermediate
metabolites (León 1999).
97
Figure 3.B.4. Disappearance ratio (%) of TEX, oxygen uptake and A255 evolution during
a representative pulse, after 4 days of cultivation.
[TEX]l0 = 25 mgC/l, [X] = 0.85 g/l, Agitation = 200 RPM, Qair = 0.5 l/min
Such studies were carried out during 30 days of process and focused on the removal of
each compound separately (Tol, EB, o-, m-, p-X).
They revealed some modifications of the bacterial behavior in the course of cultivation.
As illustrated in Figure3.B.5, from day 1 until day 20, toluene exhibited the fastest
elimination. It was followed by EB, m- and p-X, and finally o-X. (Figure 3.B.5 a and b).
Between days 22 and 25, this ranking was changed: EB reached first the lowest residual
concentration and Tol was only removed after Xylenes (Figure 3.B.5 c).
Then from day 26 until the end, the ranking was the same as at the beginning (Figure
3.B.5 d).
For all that, global rates of TEX disappearance remained unchanged (see Figure 3.B.6
below). Similar evolutions were observed thrice.
0
20
40
60
80
100
0 10 20 30 40 50 60Time (min)
Ab
sorb
an
ce (
x10
0),
[TE
X] (
%),
DO
(%
) Abs 255 nm
DO
[TEX]
98
Figure 3.B.5. Disappearance of TEX checked along 30 days of cultivation
Three points are tentatively proposed to comment the differences in the TEX removal.
o First, as the substrate loading is continuously accommodated to the biomass
growth, it is reasonable to consider the saturation and full expression of available
enzymes (Grady et al. 1996). Since global degradation and growth parameters remain
constant (see Figure 3.B.6 below), sporadic delays between supply and demand would
only tend to orientate the metabolic activity towards the EPS production.
o Second, the greater the amounts of TEX injected in the medium, the higher its
overall toxicity. Solvents are known to accumulate in and disrupt the bacterial cell
membrane thus affecting the structural and functional integrity of the cell (Sikkema et al.
1995). Therefore the major metabolic pathway could be changed. Indeed two distinct
(a) : 0 - 20 days (t = 7)
0
20
40
60
80
100
120
0 2 4 6 8 10
Time (min)
Con
cent
ratio
n (%
)
tolueneethylbenzenzem-,p-xyleneso-xylene
(b) : t = 21 days
0
20
40
60
80
100
0 2 4 6 8 10
Time (min)
Con
cent
ratio
n (%
)
(c) : 22 - 25 days (t = 23)
0
20
40
60
80
100
0 5 10Time (min)
Con
cent
ratio
n (%
) (d) : 26 - 30 days (t = 26)
0
20
40
60
80
100
0 2 4 6 8 10Time (min)
Con
cent
ratio
n (%
)
99
substrate interactions are to be considered: toluene and ethylbenzene are inhibitive
competitors, whereas xylenes transformation is necessarily cometabolic and so secondary
to toluene and/or ethylbenzene degradation (Yu et al., 2001).
o Three, the studied biodegradation is carried out by a mixed bacterial community.
Many strains are involved, and their proportion can be modified as well as the metabolic
dynamism during the cultivation process. Biomolecular techniques such as RT-PCR
(Real Time Polymerase Chain Reaction) or T-RFLP (Terminal Restriction Fragment
Length Polymorphims) would provide an idea of the bacterial composition as the process
goes along.
3.B.4.2.4. Substrate uptake rate
TEX degradation rates were investigated during the overall process.
The initial substrate uptake rate (qs,i) is calculated from the slope of the time course of the
TEX removal from the gas phase (GC analyzes). Only the first 15 minutes after injection
of TEX were considered for this calculus.
Global substrate uptake rate (qs,g) was calculated based on the total amount of TEX
removed during one pulse (by correcting for stripping).
As qs,g also took into account the degradation of both TEX and intermediate metabolites,
it was lower than the initial rate (qs,i).
Figure 3.B.6 shows the typical time course of the substrate uptake rates during the
cultivation process.
Both substrate uptake rates were maximal at the beginning of the cultivation process and
sloped down to a mean value of 0.015 gc gDW-1 h-1.
It should be noticed that the substrate uptake rate presents the rate of removal of the TEX
from the gas phase.
However it is difficult to argue if the substrate is either degraded or only absorbed inside
the bacterial cell. Therefore decreasing qs reflects a metabolic inhibition and/or an
increasing difficulty for the bacteria to incorporate the substrate. The same behavior has
100
also been reported by Bielefeldt and Stensel (1998), the longer the cultivation time, the
lower the degradation rates during the cultivation process.
A progressive deterioration of the bacterial walls is prospected to explain this evolution,
which is due to the well-known toxicological characteristics of the lipophilic substrates.
In this case, to accelerate the adaptation phase, the most efficient conditions have to be
selected. A more stable process can be achieved quickly which improves the overall
biomass productivity.
Figure 3.B.6. Evolution of the degradation rates qs during the cultivation process
3.B.5. Conclusion
In this study, SPB technique has been adopted for the cultivation of bacteria for
degradation of TEX. Obtained results indicated that relatively higher yields and
productivity could be achieved in computer controlled SPB technique compared to the
manually controlled one. Automated SPB cultivation technique benefits form an easier
maintenance coupled with improved performances. The subsequent increase of reliability
presents this process as a ready-way to produce large quantities of cell mass..
0.00
0.02
0.04
0.06
0.08
0.10
0.12
0 5 10 15 20 25 30
Time (d)
q s (
g C/g
MS h
)
dynamic qs
global qs
101
Microorganisms produced in this way can play a major role for the removal of
hydrocarbons from industrial off-gases, degradation of substitute organic solvents,
reduction in odors from wastewater plants and the food industry or even the removal of
pollutant mixtures containing chlorinated solvents.
3.B.6. Nomenclature
A255 = abs 255 nm = absorbance at 255 nm
BOD = biochemical oxygen demand
DO = dissolved oxygen
DW = dry weight
KLa = mass transfer coefficient
qs,g = global substrate uptake rate
qs,i = initial substrate uptake rate
RPM = rotation per minute
SPB = substrate pulsed batch
TEX = toluene, ethylbenzene, o-, m-, p-xylene
3.B.7. References
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Chavez-Gomez, B., 2003. Evaluation of biomass production in unleaded gasoline and
BTEX-fed batch reactors. Water Sci. Technol. 48 (8), 127-133.
Bielefeldt, A., Stensel, H., 1999. Evaluation of biodegradation kinetic testing methods
and longterm variability in biokinetics for BTEX metabolism. Wat. Res. 33 (3), 733-
740.
Bielefeldt, A., Stensel, H., 1999. Modeling competitive effects during biodegradation of
BTEX mixtures. Wat. Res. 33 (3), 707-714.
De Bont, J., 1998. Solvent-tolerant bacteria in biocatalysis. Tibtech. 16, 493-499.
102
Duenas, M., Mundate, A., Perea, A., Irastorza, A., 2003. Exopolysaccharides production
by Pedioceccus damnosus 2.6 in a semi defined medium under different growth
conditions. International Journal of Food Microbiology 2709, 1-8.
Durnaz, B., Sanin, F., 2001. Effects of carbon nitrogen ratio on the composition of
microbial extracellular polymers in activated sludge. Water Sci. Technol. 44 (1), 221-
229.
Grady, L., Smets, B., Barbeau, D., 1996. Variability in kinetic parameter estimates: a
review of possible causes and a proposed terminology. Wat. Res. 30 (3),742-748.
León, E., Seignez, C., Adler, N., Péringer, P., 1999. Growth inhibition of biomass
adapted to the degradation of toluene and xylenes in mixture in a batch reactor with
substrates supplied by pulses. Biodegradation 10, 245-250.
Liu, H., Fang, H., 2003. Influence of extracellular polymeric substances (EPS) on
flocculation, settling and dewatering of activated sludge. Crit. Rev. Env. Sci. Technol.
33 (3), 237-273.
Platt, R., Geesey, G., Davis, J., White, D., 1985. Isolation and partial chemical analysis of
firmly bound exopolysaccharides from adherent cells of fresh eater sediment bacterium.
Can J Microbiology 31, 675-680.
Raunkjaer, K., Hvitved-Jacobsen, T., Nielsen, P., 1993. Measurement of pools of protein,
carbohydrates and lipid in domestic wastewater. Wat. Res. 28, 251-262.
Seignez, C., Vuillemin, A., Adler, N., Péringer, P., 2001. A procedure for production of
adapted bacteria to degrade chlorinated aromatics. J. Hazard. Mater. B 84, 265-277.
Shim, H., Yang, S., 1999. Biodegradation of benzene, toluene, ethylbenzene, and o-
xylene by a coculture of Pseudomonas putida and Pseudomonas fluorescens
immobilized in a fibrous-bed bioreactor. J. Biotechnol. 67 (2-3, 99-112.
Sikkema, J., De Bont, J., Poolman, B., 1995. Mechanisms of solvent toxicity of
hydrocarbons. Microbiol. Rev. 269, 8022-8026.
103
Villatoro-Monzón, W.R., Mesta-Howard, A.M., Razo-Flores, E., 2003. Anaerobic
biodegradation of BTEX using Mn(IV) and Fe(III) as alternative electron acceptors.
Water Sci. Technol. 48 (6),125-131.
Yu, H., Kim, B., Rittmann, B., 2001. A two-step model for the kinetics of BTX
degradation and intermediate formation by Pseudomonas putida F1. Biodegradation 12
(6), 465-475.
Yu, H., Kim, B., Rittmann, B., 2001. The roles of intermediates in biodegradation of
benzene, toluene, and p-xylene by Pseudomonas putida F1. Biodegradation 12 (6),
455-463.
105
CHAPTER 4.
Photocatalytic Degradation of p-halophenols in TiO2 Aqueous Suspensions: Halogen Effect on Removal Rate, Aromatic Intermediates and Toxicity Variations
4.1. Abstract
The influence of the halogen upon the TiO2 photocatalytic degradation of p-halophenols
in water has been investigated in this chapter. Phenol was used as the reference
compound.
Compared with its value for phenol, the apparent first-order rate constant of removal, k,
was slightly but significantly higher for p-fluorophenol and p-chlorophenol, and slightly
but significantly lower for p-bromophenol. For p-iodophenol, k was about half that of
phenol. The relative values confirm that k is roughly correlated to the Hammett constant;
this constant reflects the electron density on the aromatic ring and, accordingly, the
reactivity towards electrophilic species generated by UV-irradiation of TiO2.
All compounds were found to be poorly adsorbed on TiO2. Accordingly, k was not
related to the differences observed in the very low adsorbed amounts.
The detected aromatic intermediate products included hydroquinone (HQ), benzoquinone
(BQ) and various halodihydroxybenzenes. HQ, BQ, 4-chloro-1,2 (and 1,3)-
dihydroxybenzenes and 4-bromo-1,3-dihydroxybenzene were quantified. Mechanisms are
tentatively suggested to interpret the differences in the degradation pathways of the p-
halophenols. The organic intermediate products accounted for only a few percents of the
total carbon during the degradation.
The toxicity (1/EC50) measured by the MicrotoxTM test almost did not vary in the course
of the degradation of phenol, p-chlorophenol and p-bromophenol until complete removal
Accepted for publication in Journal of Environmental Science and Health, Part A
106
of these compounds. By contrast, the value of 1/EC50 was multiplied by ca. 2.5 when ca.
45% of p-iodophenol had been removed; concentrations of BQ higher than with the other
p-halophenols are tentatively suggested to be at the origin of this increase. Interpretation
of a surprising substantial increase in the toxicity value when the removal of p-
fluorophenol increased from 80 to 95% requires further investigation.
4.2. Introduction
The photocatalytic degradation of chloroaromatic pollutants in water has been widely
studied (Blake, 1999; Pichat, 2003; Fujishima et al., 1999). In particular, it has been
shown that benzenes substituted, inter alia, by chlorine can be dechlorinated.
However, investigations regarding benzene derivatives substituted by other halogens are
rarer (Stafford et al., 1996), principally because fluorinated, brominated and iodinated
compounds are less often employed. Nevertheless, studying their photocatalytic
behaviour is also of interest to better assess the capabilities of this purification method
and conceivably to provide insights into the degradation mechanisms.
In this chapter, the photocatalytic degradation of p-halophenols in TiO2 aqueous
suspensions is reported.
The study includes: the extent of adsorption of these compounds, the disappearance rates,
the analyses of the primary aromatic intermediate products, and the variations in total
organic carbon and toxicity (evaluated by the MicrotoxTM test) during the degradation.
The purposes were to determine how the nature of the halogen affects these
characteristics and whether the potential differences depend on the experimental
conditions.
In particular, it was wished to further document the effects on the removal rates (i) of
adsorption for these poorly adsorbed compounds, and (ii) of electron density on the
aromatic ring, which varies with the halogen substituent.
It was also expected that changing the halogen might provide additional information on
the basic photocatalytic degradation mechanisms, which are still controversial.
Another objective was to assess to what extent the variations in toxicity during the
photocatalytic treatment differ for structurally closely related pollutants.
107
4.3. Materials and methods
4.3.1. Materials
All organic compounds were purchased from Aldrich, except 4-chlorophenol obtained
from TCI. They were used without further purification (purity > 98%). The photocatalyst
was TiO2 Degussa P25 (about 80 % anatase, 20% rutile, 50 m2 g-1, non porous). To check
the photocatalyst effect on the ranking of the p-halophenols degradation rates,
Millennium PC 50 (100% anatase, 50 m2 g-1, average pore diameter: 20 nm) was used.
4.3.2. Adsorption isotherms in the dark
The batch adsorption equilibrium experiments were conducted at natural pH (ca. 6.5).
Preliminary experiments, performed during 24 h under constant magnetic stirring,
showed that the equilibrium was reached within 1 h. The equilibrium concentrations, Ceq,
were determined by liquid chromatography after centrifugation and filtration (0.45 µm
Millipore filters) of the suspension.
The number nM of moles adsorbed per gram of TiO2 was calculated as follows:
W
CVnM
∆⋅= , (Eq 1.1)
where ∆C was the difference between the initial concentration (up to 30 mmol L-1) and
Ceq, V was the suspension volume (0.02 l), and W was the TiO2 mass (0.2 g).
4.3.3. Photocatalytic degradation experiments
Unless otherwise indicated, a Philips HPK 125 W high-pressure mercury lamp was used.
Its output was filtered by a 2.2 cm thick circulating-water cuvette to absorb the infrared
wavelengths and a 340 nm cut-off filter (Corning 0.52) to avoid direct photolysis of the
organic pollutants. The radiant flux thus entering the reactor was 47±4 mW cm-2 (as
measured with a UDT 21 A powermeter). The magnetically stirred, aqueous suspensions
were UV-irradiated in a cylindrical flask (total volume: ca. 90 ml), opened to air, with a
bottom optical window whose surface was approximately 11 cm2. The solutions were
prepared with deionized, doubly-distilled water. The volume of aqueous phenol or p-
108
halophenol solution (0.155 mmol l-1) introduced into the photoreactor was 20 ml. The
minimal amount of TiO2 necessary to have no UV radiant flux above the suspension was
0.5 g l-1. That meant that the UV photons entering the reactor were absorbed, scattered or
reflected mainly by the slurry and secondarily by the reactor. The degradation was carried
out at room temperature and at natural pH (ca. 6.5). The aerated suspension was first
stirred in the dark for 60 min to reach equilibrated adsorption.
To check the effects of the experimental conditions on the ranking of the degradation
rates for the four p-halophenols and phenol, another type of reactor and a xenon lamp
were also used. The total radiant flux, measured with a YSI Corporation powermeter, was
80 mW cm-2. About 0.5% of the photons were emitted at wavelengths shorter than 300
nm and about 7% between 300 and 400 nm. This lamp was part of a Hanau Suntest
Simulator. A 290 nm cut-off filter was used. The photocatalytic experiments were
performed in open-to-air Pyrex flasks placed in the Suntest Simulator and containing 40
ml of suspension.
4.3.4. Analytical methods
The high performance liquid chromatography (HPLC) system comprised a LDC
constametric 3000 isocratic pump and a LDC Spectro Monitor D UV detector adjusted at
225 nm. A reverse-phase column, 25 cm long, 4.6 mm i.d., packed with Spherisorb 5
ODS-2 was used. The mobile phase was composed of methanol (40 v/v %) and deionized
doubly distilled water at pH 3 (adjusted with orthophosphoric acid).
Under these conditions, the retention times in min-1 were approximately: 10 (p-
fluorophenol); 21 (p-chlorophenol); 26 (p-bromophenol); 43 (p-iodophenol); 3
(hydroquinone); 5 (benzoquinone); 3 (1,2,4-trihydroxybenzene); 7 (4-chloro-1,3-
dihydroxybenzene); 11.5 (4-chloro-1,2-dihydroxybenzene); 8.5 (4-bromo-1,3-
dihydroxybenzene).
Experience shows that retention times vary with trademarks, previous uses of the column
and other settings of the HPLC apparatus.
Consequently, the aforementioned values must be regarded as indicative, and differences
should not be disquieting for someone attempting to copy the method.
109
Identification of the intermediate organic products was performed both by comparing
their UV spectrum to those of commercial compounds by use of both a Varian-9065
Polychrom diode-array HPLC detector and by GC-MS (HP 5890 and 5971 A).
For non-commercialised compounds, samples were analysed with liquid chromatography
/ mass spectrometry (LC-MS: HP Series 1100; Inersil ODS-2 column, 10 cm long, 3 mm
i.d.). The mobile phase was a mixture of methanol and deionized, doubly distilled water,
the v/v percentage of methanol varying from 10 to 90 over 30 min.
Dissolved Organic Carbon measurements were performed using a TOC analyzer
(Shimadzu, model 5050A) equipped with an ASI automatic sample injector. A potassium
phthalate solution was utilised for calibration. The precision and limit of detection of the
apparatus were 1 and 0.2 ppm, respectively.
For analysing the halide anions, a Sarasep AN1 anion exchange column (Touzart, 10 cm
long, 4.6 mm i.d.) with a solid phase chemical suppressor cartridge (Alltech) was
employed. The mobile phase was a mixture of sodium hydrogenocarbonate (1.7 mmol l-1)
and sodium carbonate (1.8 mmol l-1) in deionized, doubly distilled water. The detector
was a Waters 431 apparatus. pH values were measured with a Iono Porcessor II
(Tacussel).
The acute toxicity was assessed using a MicrotoxTM Model 500 Analyzer. This analyzer
is based on the use of a luminescent bacterium, viz. the strain Vibrio fischeri NRRL B-
11177. When properly grown, luminescent bacteria produce light as a by-product of their
cellular respiration. Any inhibition of cellular activity results in a decreased rate of
respiration and a corresponding decrease in the rate of luminescence. The toxicity is
determined by measuring, after 5 and 15 minutes of exposure, the concentration at which
50% of the light is lost because of the toxicity of the solution examined (EC50). The
Toxic Unit (TU) is used to express the data, as proposed by U.S. EPA (Lankford and
Eckenfelder, 1990): TU = 100/EC50.
In all cases, the analyses were at least duplicated. The standard deviation was less than
5%. The results reported in this study correspond to the average value of the
measurements.
110
4.4. Results and discussion
4.4.1. Adsorption of phenol and p-halophenols on TiO2
Figure 4.1 shows the results obtained for the extent of adsorption of p-halophenols on
TiO2 as a function of Ceq, the concentration in bulk solution at adsorption/desorption
equilibrium in the dark.
The number nM of organic molecules adsorbed per gram of TiO2 can be derived from the
following relationships (Cunningham and Srijaranai, 1994; Chen and Ray, 1999; Al-
Sayyed et al., 1991):
maxmax. M
eq
MM
eq
n
C
nKn
C+= 1 Eq 4.2
eqsp
oA
sp
oA
M
eqC
A
N
KA
N
n
C.
.
.
. σσ+= Eq 4.3
where K is the adsorption equilibrium constant for the compound considered, nMmax is the
maximum value of nM, σ° is the average area occupied by one adsorbed organic
molecule, Asp is the TiO2 surface area, and NA is the Avogadro’s number.
The plots of Ceq/nM versus Ceq are shown for the four p-halophenols using the right-hand
scale in Figure 4.1.
The values of nMmax and σ° are calculated from the slope and ordinate at the origin of the
straight lines in Figure 4.1 according to Equations 2 and 3. These values are listed in
Table 4.1.
The number of molecules adsorbed by surface area unit of catalyst, viz. (1/σ°) is in the
order:
p-iodophenol > phenol > p-chlorophenol > p-bromophenol > p-fluorophenol
111
Figure 4.1. Room temperature isotherms of adsorption on TiO2 for phenol and p-
halophenols. For the meaning of nM and Ceq, see equations (2-3). The thicker curves refer
to the variations of nM.
phenol
y = 17.177x + 535.58R2 = 0.9515
0
0.1
0.2
0.3
0.4
0 10 20 30 40
Ceq*10-4 / mol l-1
nM*1
0-5 /
mo
l g-1
0
500
1000
1500
Ce
q/n
M
fluorophenol
y = 55.852x - 513.94R2 = 0.9896
0
0.1
0.2
0.3
0.4
0 5 10 15 20 25 30
Ceq*10-4 / mol l-1
nM*1
0-5 /
mo
l g-1
0
500
1000
1500
Ceq
/nM
chlorophenol
y = 26.313x + 177.07R2 = 0.982
0
0.1
0.2
0.3
0.4
0 5 10 15 20 25 30
Ceq*10-4 / mol l-1
nM*1
0-5 /
mo
l g-1
0
500
1000
1500
Ce
q/n
Mbromophenol
y = 36.74x + 173.27R2 = 0.99
0
0.1
0.2
0.3
0.4
0 5 10 15 20 25 30
Ceq*10-4 / mol l-1
nM*1
0-5 /
mo
l g-1
0
500
1000
1500C
eq/n
M
iodophenol
y = 12.97x + 191.34R2 = 0.75
0
0.1
0.2
0.3
0.4
0 5 10 15 20 25 30
Ceq*10-4 / mol l-1
nM*1
0-5 /
mo
l g-1
0
500
1000
1500
Ceq
/nM
112
For p-chlorophenol, it agrees with previous reports (Al-Sayyed et al., 1991; D'Oliveira et
al., 1990). The values are lower than those measured for p-chloro and p-
fluoromethoxybenzenes (anisoles), viz. about 0.2 molecules per nm2 (Amalric et al.,
1996).
Considering that the geometric area of one molecule of p-halophenol is ca. 0.5 nm2 (20 to
25% larger than benzene, i.e. 0.4 nm2; Gregg and Sing, 1967), about two molecules
would be needed to cover 1 nm2 of catalyst. In other words, the surface coverage
measured was comprised between 1.0 and 4.5%.
Low coverage is expected because pollutant molecules need to break the three-
dimensional network of hydrogen-bound water molecules close to the TiO2 surface and to
reorganize the water molecules around them. Therefore an unfavorable change in entropy
occurs.
Also, if it is assumed that p-halophenol molecules are hydrogen-bound to the surface
hydroxyl groups and that the maximum concentration is on the order of 5 OH groups per
nm2 of TiO2 (Boehm, 1966), at most 5 molecules of halophenol per nm2 could be
adsorbed. With respect to this maximum, the values measured would correspond to only
0.2 to 0.9 %.
Table 4.1. Measured extents of adsorption (Conditions: as for Figure 4.1), molecular
characteristics (Handbook of Chemistry and Physics, 1992) and measured values of k for
phenol and p-halophenols. (Conditions: as for Figure 4.2)
nMmax (µmol g-1) molecules adsorbed per nm2
of TiO2 (1/σo)x 102
C-H and C-F bond energy (kJ mol-1)
Hammet constant
σ
k
(h-1)
phenol 5.8 7.0 368 (C-H) 0 1.98
p-fluorophenol 1.8 2.2 523 (C-F) 0.15 2.17
p-chlorophenol 3.8 4.6 397 (C-Cl) 0.24 2.09
p-bromophenol 2.8 3.5 335 (C-Br) 0.26 1.76
p-iodophenol 7.7 9.3 272 (C-I) 0.28 0.88
113
4.4.2. Photocatalytic removal rates of phenol and p-halophenols
4.4.2.1. Results
The influence of the halogen substituent on the photocatalytic degradation of the p-
halophenols was investigated by comparing their apparent rate constants of disappearance
(k). In all cases, first-order kinetics was observed for up to 15 minutes, with linearization
coefficients generally > 0.99 (see inset in Figure 4.2). The first 15 min have been used to
focus on rates of elimination of the parent compound and to minimize competition with
the removal of intermediate products.
The values of k (Table 4.1) are in decreasing order as follows:
p-fluorophenol ≈ p-chlorophenol ≈ phenol > p-bromophenol > p-iodophenol.
Figure 4.2. Kinetics of the photocatalytic disappearance of phenol (diamonds), p-
fluorophenol (hyphens), p-chlorophenol (circles), p-bromophenol (squares), and p-
iodophenol (triangles). The inset shows linear transforms. C is the concentration of the
pollutant considered. Subscipt 0 refers to the initial conditions.
Conditions: 20 ml; 0.155 mmol l-1; TiO2 Degussa P25 (0.5 g l-1); 47 mW cm-2
0
20
40
60
80
100
120
140
160
0 50 100 150
t (min)
C (
µmol
l-1)
0
0.2
0.4
0.6
0 5 10 15t (min)
ln (
Co/
C)
114
This ranking is similar to that previously reported (O'Shea and Cardona, 1994; Parra et
al., 2000).
This order did not change when TiO2 Millennium PC 50 was used in place of TiO2
Degussa P25 under otherwise identical conditions. Note that the two samples have the
same surface area but PC 50 is porous, whereas P25 is not. The k values corresponding to
P25 were about 1.6 times higher on the average.
When another reactor and lamp were employed (see Experimental section), the k ranking
was still the same when increasing, under otherwise identical conditions, either the P25
concentration by a factor of 2 or the pollutant initial concentration by a factor of 2.5.
However, the ratio of k for each halophenol with respect to k for phenol depended to
some extent on the conditions.
This shows that attempts to correlate results obtained under various conditions have
limitations (Serpone et al., 1996).
4.4.2.2. Correlation between the photocatalytic degradability
and the adsorbed amounts of p-halophenols
It is a priori reasonable to assume that the closer the pollutant to the photocatalyst
surface, the higher the probability of reaction with holes or other oxidizing species
formed on that surface.
In fact, p-iodophenol has the lowest k and the highest, though very low, number of
molecules adsorbed per nm2 of TiO2 (Table 4.1).
And comparison of the other values of k and nM for the other p-halophenols clearly shows
that there is no relationship.
Indeed, all p-halophenols are poorly adsorbed as aforementioned.
4.4.2.3. Correlation between the photocatalytic degradability
and the Hammett constant of p-halophenols
The Hammett constant σ represents the electron density over the aromatic ring. The
higher σ the lower the electron density. In agreement with previous reports (O'Shea and
115
Cardona, 1994; Parra et al., 2000; D'Oliveira et al, 1993), a rough negative correlation
was found between σ and k with a R2 coefficient comprised between 0.57 (case of the k
listed in Table 4.1) and 0.95 depending on the conditions (type of reactor, lamp and TiO2,
etc.).
This correlation suggests an attack by electrophilic species, i.e. hydroxyl radical or hole
(Walling et al., 1978).
4.4.3. Aromatic intermediate products of degradation
The aromatic intermediate products were identified by LC-MS using commercially
available standards, viz. hydroquinone (HQ), benzoquinone (BQ), 1,2,4-
trihydroxybenzene, 4-chloro-1,2 (and 1,3)-dihydroxybenzenes, and 4-bromo-1,3-
dihydroxybenzene.
Besides, the formation of another bromodihydroxybenzene, a fluorodihydroxybenzene
and a dihydroxyiodobenzene was inferred from the mass spectra by comparison with the
spectra of the commercialized halodihydroxybenzenes.
The additional brominated derivative was assumed to be 4-bromo-1,2-dihydroxybenzene
as a result of the directive effect of the OH group of 4-bromophenol towards the ortho
position for an electrophilic attack (Scheme 4.1).
As the positional selectivity of the F atom is the strongest and that of the I atom the
weakest, 4-fluoro-1,3-dihydroxybenzene and 1,2-dihydroxy-4-iodo-benzene were likely
to be the other halogenated derivatives detected.
OH
OH
X
OHOH
X
OH
OO
O2
- HO2.
OH
X
CH
OH
X
OH
or h+;
H2O;-H+
+
or h+;
Scheme 4.1. Suggested mechanism for the formation of 4-halo-1,2-dihydroxybenzene
116
Standards allowed us to quantify the amounts of the two halogenated aromatic
intermediate products of 4-chlorophenol.
Figure 4.3 shows that 4-chloro-1,3-dihydroxybenzene reached a higher percentage than
4-chloro-1,2-dihydroxybenzene in the course of the degradation. In principle, the
directive effect of the OH group should prevail over that of the Cl atom, therefore the
percentages should have been the opposite of those observed. Lower amounts of 4-
chloro-1,2-dihydroxybenzene may be explained by either a higher extent of adsorption of
this compound on TiO2 (Kads = 0.13 l µmol-1; Pichat, 2003), or a higher degradation rate.
The first hypothesis seems more likely (Pichat, 2003).
The concentration of 1,2,4-trihydroxybenzene remained very low in accordance with the
well-known photocatalytic instability of this compound.
The sum of the concentrations of HQ and BQ was higher than the concentrations of the
aromatic halogenated intermediate products (Figures 4.3 and 4.4). Moreover, the
concentration of BQ was higher than that of HQ in the case of p-iodophenol, whereas it
was the opposite for the other p-halophenols and phenol (Figure 4.4). In particular, for p-
fluorophenol the maximum concentration of BQ reached only 0.3 µmol l-1, that is,
approximately 60 times less than the maximum concentration of HQ.
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0 50 100
Conversion (%)
Inte
rmed
iate
s (%
)
Figure 4.3. Percentages of 4-chloro-1,3-dihydroxybenzene (solid circles) and 4-chloro-
1,2-dihydroxybenzene (open circles) formed with respect to the initial pollutant
concentration versus the percentage of pollutant degraded. Conditions: as for Figure 4.2.
117
Figure 4.4. Percentages of BQ, HQ and BQ + HQ (with respect to the initial pollutant
concentration) versus the removed percentage of phenol (diamonds), p-fluorophenol
(hyphens), p-chlorophenol (circles), p-bromophenol (squares), and p-iodophenol
(triangles). Conditions: as for Figure 4.2
Scheme 4.1 shows the mechanism usually proposed for hydroxylation of a benzene
nucleus via the formation of a hydroxyhexadienyl radical. It accounts for the formation of
the various 4-halodihydroxybenzenes.
In the case of phenol, formation of HQ can occur through a hydroxycyclohexadienyl
radical if the attack occurs at position 4.
Replacing attack by an OH radical by attack by a hole, addition of H2O and elimination
of H+ would yield to the same initial organic radical. In the case of the p-halophenols, the
C atom in position 4 is not prone to be attacked by an electrophilic species because of the
0
2
4
6
8
10
12
14
0 20 40 60 80 100
Hyd
roqu
inon
e (%
)
0
0.5
1
1.5
2
2.5
0 20 40 60 80 100
Ben
zoqu
inon
e (
%)
0
5
10
15
0 20 40 60 80 100Conversion (%)
Ben
zo+hy
droq
uino
ne (
%)
118
electron-withdrawing effect of the halogen X. Consequently, formation of HQ via the
sequence of reactions depicted in Scheme 4.2a is unlikely.
Scheme 4.2. Possible degradation pathways of p-halophenols, during TiO2
photocatalysis, leading to either hydroquinone (HQ) (a and b) or benzoquinone (BQ) (d),
and HQ oxidation into BQ (c). For discussion, see text.
To account for the formation of HQ from p-halophenols and for the differences induced
by the nature of the halogen, we tentatively suggest a reductive attack at position 4 by a
OH
X
OHC
XOH
OH
C
OH
O
- HX
OH
OH
OH
X
C
OH OH
OO
OH
OH
- X-
- R.+ RH
e-O2
- HO2.
+ H2O
OH
X
OH
O
X
C
O
X
O
O
- H2O
HO2.
- HXO
OH
OH OH
O O
O
O2
OH
OH
.+
h+
- H+ - HO2.
(a)
(b)
(c)
(d)+ / O2.-
/XO-
119
photoproduced electron with the release of a X- anion. As shown in Scheme 4.2b, the
resulting aryl radical, after O2 addition, would yield HQ, H2O being used as a H source.
The higher the electronegativity of X, the easier the reductive attack. That is in agreement
with the concentration of HQ observed during the degradation, the highest concentration
corresponding to p-fluorophenol and the lowest to p-iodophenol (Figure 4.4).
The mechanism suggested implies the cleavage of the C-X bond whose strength depends
on the electronegativity of X. On the basis of this only argument, p-iodophenol should
yield the highest amount of HQ. But the C-X bond strength intervenes in a second step,
the primary step being the reductive attack which is favored by the partial positive charge
on the C atom and, accordingly, by the electronegativity of X (inductive effect).
BQ can be formed by oxidation of HQ (Scheme 4.2c). If this oxidation was the only
pathway, the ratio between BQ and HQ concentrations would be independent of the p-
halophenol. As this is not the case, a mechanism implying position 4 on the aromatic ring
must be involved.
The mechanism tentatively suggested (Scheme 4.2d) includes first the removal of the H
atom of the phenolic OH group, second the use of HO2°/O2°- as a source of O with the
release of HXO/XO- (in the case of phenol: H2O/OH-), at least for X = I, Br, Cl. Note that
this mechanism implies a reaction between two radicals, which diminishes its probability
to occur; however, HO2°/O2°- are relatively abundant radicals during UV irradiation of
TiO2 aqueous suspensions in the presence of air. As the cleavage of the C-X bond occurs,
the lower this bond strength (Table 4.1), the higher the reaction rate.
This is in agreement with the results (Figure 4.2 and Table 4.1). In addition, hypo-iodous,
-bromous and -chlorous acids and the corresponding conjugate base anions are stable
compounds, which might also explain a markedly higher rate of formation of BQ from p-
iodo, bromo, and chlorophenol compared with p-fluorophenol (Figure 4.4).
The fluorous equivalents are highly unstable and their decomposition to HF/F- and O2
makes a priori the mechanism less likely to take place.
120
4.4.4. Toxicity
According to the MicrotoxTM tests, the toxicity ranking of the initial solutions from the
less to the most toxic was:
phenol < p-fluorophenol < p-chlorophenol < p-iodophenol < p-bromophenol
(see Figure 4.5).
This tendency is in agreement with previous results (McCloskey et al., 1996; Cronin and
Schultz, 1996), except for p-bromophenol, which was found less toxic than p-iodophenol.
Figure 4.5. Evolution of the acute toxicity during the photocatalytic degradation of
phenol (diamonds), p-fluorophenol (hyphens), p-chlorophenol (circles), p-bromophenol
(squares), and p-iodophenol (triangles). Top: Kinetics. Bottom: Variation as a function of
the percentage of pollutant removed. Conditions: as for Figure 4.2.
0
200
400
600
800
1000
1200
1400
1600
1800
0 15 30 45 60t (min)
Tox
icit
y (T
U)
0
200
400
600
800
1000
1200
1400
1600
1800
0 25 50 75 100Conversion (%)
Tox
icit
y (T
U)
121
During the photocatalytic degradation of phenol and p-halophenols, distinct toxicity
variations were observed depending on the initial compound. For phenol and p-
chlorophenol, the toxicity increased slightly and for p-bromophenol it decreased slightly.
A progressive rise was measured during the degradation of p-iodophenol, so that the
MicrotoxTM toxicity value was ca. 2.5 times higher when ca. 45% of p-iodophenol had
been eliminated. A very pronounced increase in toxicity was observed when the removal
of p-fluorophenol increased from 80 to 95%.
Though toxic (Corti and Snyder, 1998; De Caprio, 1999), HQ had not a significant effect
on the overall toxicity. Indeed, HQ attained its highest, though very low concentrations in
the course of the degradation of phenol and p-fluorophenol (Figure 4.4), while the
measured toxicity remained the weakest of all the solutions examined (Figure 4.5).
The concentration of BQ, a compound more toxic than HQ (Den Besten et al., 1994;
Colinas and Burkart, 1994), might explain the increase in the toxicity of the p-
iodophenol-containing solution during degradation.
But this suggestion is only tentative as the percentage of BQ, with respect to the quantity
of p-iodophenol eliminated, remained very low (Figure 4.4).
It has recently been reported that the cytotoxicity of iodoacetic acid is much higher than
that of chloroacetic acid (Plewa et al., 2004). Accordingly, the observed increase in
toxicity might be due to traces of I-containing organic intermediate products.
In the case of the p-fluorophenol degradation, the sudden increase in toxicity (Figure 4. 5)
corresponded to a decrease in the concentration of F- anions. Further work would be
needed to determine whether these phenomena are coincidental or related, e.g. the
formation of a SiF62-complex might be assumed to occur as a result of an attack of the
reactor glass by HF.
This discussion is meant to be only suggestive. Interpreting the toxicity variations would
imply a complete knowledge of the nature and toxicity - on the bacteria employed - of the
degradation products, which cannot be achieved. However, it is known that total
mineralization can be reached by use of sufficiently prolonged UV-irradiation.
122
Figure 4.6. Percentages of the relative carbon forms in the TiO2 suspensions during the photocatalytic degradation of p-halophenols (p-halophenols in gray, quantified
intermediate products in black, CO2 in striped and the difference in white). Conditions: as for Figure 4.2.
0
20
40
60
80
100
0 5 10 15 30 60 120 180t (min)
perc
enta
ges
iodophenol
0
20
40
60
80
100
0 5 10 15 30 60 120 180t (min)
perc
enta
ges
bromophenol
0
20
40
60
80
100
0 5 10 15 30 60
t (min)
perc
enta
ges
chlorophenol
0
20
40
60
80
100
0 5 10 15 30 60
t (min)
perc
enta
ges
fluorophenol
0
20
40
60
80
100
0 5 10 15 30 60t (min)
perc
enta
ges
phenol
123
4.4.5. Carbon balance
Figure 4.6 shows the kinetic evolutions in the various forms of carbon. The time scale
shown coincides with the total or almost total disappearance of the original compound.
As pointed out in several papers dealing with the photocatalytic degradation of p-
chlorophenol or its isomers (Chen and Ray, 1999; Al-Sayyed et al., 1991; D'Oliveira et
al., 1990):
o the amount of aromatic intermediate products corresponded to only a few % of
the initial amount of the original pollutant (less than 5% of TOC in the present case all
along the degradation), and
o these aromatics disappeared at about the same time as the original compound.
The present study indicates that these features were similar for the other p-halophenols.
4.5. Conclusion
Beyond confirming the importance of the overall electron density on the ring for the
photocatalytic removal rate of substituted benzenes in water, comparison of the primary
aromatic products of p-halophenols has allowed us to suggest degradation mechanisms
whose relative significance depends on the halogen.
Also, even though the total quantity of these aromatic intermediate products was found to
be similar regardless of the halogen, at equal removal percentage of p-halophenol the
variations in toxicity markedly differed for solutions that initially contained p-
iodophenol.
This observation illustrates that halogen-induced differences in mechanisms and
reactivity can be significant for the detoxification of waters polluted by these types of
compounds.
124
4.6. References
Al-Sayyed, G., D'Oliveira, J.-C., Pichat, P., 1991. Semiconductor-sensitised photodegradation of 4-chlorophenol in water. J. Photochem. Photobiol., A : Chem. 58, 99-114.
Amalric, L., Guillard, C., Blancbrude, E., Pichat, P., 1996. Correlation between the photocatalytic degradability over TiO2 in water of meta and para substituted methoxybenzenes and their electron density, hydrophobicity and polarizability properties. Water Res. 30, 1137-1142.
Blake, D., 1999. Bibliography of work on the heterogeneous photocatalytic removal of hazardous compounds from water and air, Update number 3 to January 1999, NREL/TP-570- 26797, National Technical Information Service (NTIS), Springfield, VA. (http://www.nrel.gov/docs/fy99osti/26797.pdf).
Boehm, H.P., 1966. Chemical identification of surface groups. Adv. Catal. 16, 179.
Chen, D., Ray, A.K., 1999. Photocatalytic kinetics of phenol and its derivatives over UV irradiated TiO2. Appl. Catal., B: Environ. 23, 143-157.
Colinas, R., Burkart, P., 1994. In vitro effects of hydroquinone, benzoquinone, and doxorubicin on mouse and human bone marrow cells at physiological oxygen partial pressure. Toxicol. Appl. Pharmacol. 129, 95-102.
Corti, M., Snyder, C., 1998. Gender- and Age-Specific Cytotoxic Susceptibility to Benzene Metabolites in Vitro. Toxicol. Sci. 41, 42-48.
Cronin, M., Schultz, T., 1996. Structure-toxicity relationships for phenols to Tetrahymena pyriformi. Chemosphere. 32, 1453-1468.
Cunningham, J., Srijaranai, S., 1994. Adsorption of model pollutants onto TiO2 particles in relation to photoremediation of contaminated water. In Aquatic and surface photochemistry, G.R. Helz, R.G. Zepp, D.G., Crosby, Eds., Lewis Publishers: Boca Raton, pp. 317-348.
De Caprio, A., 1999. The toxicology of hydroquinone - Relevance to occupational and environmental exposure. Crit. Rev. Toxicol. 29, 283-330.
Den Besten, C., Brouwer, A., Rietjens, I., Van Bladeren, P., 1994. Biotransformation and toxicity of halogenated benzenes. J. Hum. Exp. Toxicol. 13, 866-875.
D'Oliveira, J.-C., Al-Sayyed G., Pichat, P., 1990. Photodegradation of 2 - and 3 - chlorophenol in TiO2 aqueous suspensions. Environ. Sci. Technol. 24, 990-996.
D'Oliveira, J.C., Minero, C., Pelizzetti, E., Pichat, P., 1993. Photodegradation of dichlorophenols and trichlorophenols in TiO2 aqueous suspensions - kinetic effects of the positions of the Cl atoms and identification of the intermediates. J. Photochem. Photobiol. A: Chem. 72, 261-267.
Fujishima, A., Hashimoto, K., Watanabe, T., 1999. In TiO2 Photocatalysis; BKC, Tokyo.
125
Gregg, J., Sing, S., 1967. In Adsorption, Surface Area and Porosity; Academic Press, London, pp. 80.
Handbook of Chemistry and Physics, 72nd Ed., 1991-1992. D. Lide, Ed.; CRC Press, Boston.
Lankford, P., Eckenfelder Jr., W., 1990. Toxicity Reduction in Industrial Effluents. Van Vostrand Reinhold, New York, USA.
McCloskey, J., Newman, M., Clark, S., 1996. Predicting the relative toxicity of metal ions using ion characteristics: Microtox® bioluminescence assay. Environ. Toxicol. Chem. 15, 1730-1737.
O'Shea, K.E., Cardona, C., 1994. Hammett Study on the TiO2-Catalyzed Photooxidation of Para-Substituted Phenols. A Kinetic and Mechanistic Analysis. J. Org. Chem. 59, 5005-5509.
Parra, S., Olivero, J., Pacheco, L., Pulgarin, C., 2000. Structural properties and photoreactivity relationships of substituted phenols in TiO2 suspensions. Appl. Catal, B: Environ. 27, 153-168.
Pichat, P., 2003. Photocatalytic degradation of pollutants in water and air: basic concepts and applications. In Chemical Degradation Methods for Wastes and Pollutants: Environmental and Industrial Applications, M.A. Tarr, Ed., Marcel Dekker Inc., New York, pp. 77-119.
Plewa, M.J., Wagner, E.D., Richardson, S.D., Thruston, Jr., A.D., Woo, Y.T.; Bruce McKague, A., 2004. Chemical and Biological Characterization of Newly Discovered Iodoacid Drinking Water Disinfection Byproducts. Environ. Sci. Technol. 30, 4713-4722.
Serpone, N., Sauvé, G., Koch, R., Tahiri, H., Pichat, P., Piccini, P., Pelizzetti, E., Hidaka, H., 1996. Standardization protocol of process efficiencies and activation parameters in heterogeneous photocatalysis: relative photonic efficiencies ξr. J Photochem. Photobiol., A. Chem. 94, 191-203.
Stafford, U., Gray, K.A., Kamat, P.V., 1996. Photocatalytic degradation of organic contaminants: Halophenols and related model compounds. Heterogen. Chem. Rev. 3, 77-104.
Walling, C., Camaioni, D., Kim, S.S., 1978. Aromatic hydroxylation by peroxydisulfate. J. Am. Chem. Soc. 100, 4814-4818.
126
127
CHAPTER 5.
Integrated Photocatalytic-Biological Process for Treatment of Pesticides
PART A.
EVALUATING MICROTOX© AS A TOOL FOR BIODEGRADABILITY ASSESSMENT OF PARTIALLY TREATED SOLUTIONS OF
PESTICIDES USING Fe3+ AND TiO2 SOLAR PHOTO-ASSISTED PROCESSES
5.A.1. Abstract
To shorten photo-treatment time is of major concern for the cost and energy benefits of
the xenobiotics degradation performed by photocatalytic processes. Using photo-Fenton
and TiO2 phototreatments, partially photodegraded solutions of 6 separate pesticides
(alachlor, atrazine, chlorfenvinphos, diuron, isoproturon and pentachlorophenol) were
tested for biocompatibility, which was evaluated according to the Zahn-Wellens
procedure. This study investigated if Microtox© could be considered as a suitable global
indicator capable of giving information on the evolution of biocompatibility of the water
solution contaminated with organic pollutants during the phototreatment in order to
promote biotreatment. The obtained results demonstrated that biodegradability increased
significantly after short photo-Fenton treatment times for alachlor, diuron and
pentachlorophenol. Uncertain results were obtained with atrazine and isoproturon.
Microtox© acute toxicity testing was shown to correctly represent dynamics and
efficiency of photo-treatment.
Submitted to Ecotoxicological and Environmental Safety
128
5.A.2. Introduction
For mixed household and industrial wastewater treatment plants, the contribution from
industry must be managed properly in order to avoid toxic or persistent substances end up
in the effluent or sludge. (Sarakinos et al, 2000; Chen et al., 1999; Grüttner, 1997; Grau
and Da-Rin, 1997)
The magnitude of the potential risk of these toxicants on environment and human health
is widely evaluated through biological monitoring. (Mantis et al., 2005; Pessala et al.,
2004; Ma et al., 2005) In case of pesticides, the substances differ from most other
industrial organic chemicals in that they are produced with the explicit intention of
exerting toxic effects on one or more target organisms. Unfortunately, they can adversely
affect other living beings that happen to be nearby, because these compounds are brought
into the environment through various possibilities (Jones et al., 2004; Fent et al., 2006;
Aksu, 2005; Welp and Brümmer, 1999).
Advanced Oxidation Processes (AOPs) have been proposed as an alternative for the
treatment of biorecalcitrant wastewater. Many studies have concentrated on this goal
(Legrini et al., 1993; Andreozzi et al., 1999; Espulgas et al., 2002; Pera-Titus et al.,
2004). They pointed out that these processes, while making use of different reacting
systems, are all characterized by the same chemical feature: production of strong oxidant
species, such as hydroxyl radicals, which are able to oxidize almost all organic pollutants
(Valavanidis et al., 2006). The photo-assisted catalysis involves the irradiation of a semi-
conductor like TiO2 with UV light at a wavelength below 387 nm, generating electron-
hole pair on the catalyst surface and inducing the formation of radical reduced oxygen
species. These radicals are produced in situ and are highly reactive and unselective
oxidants. These OH radicals can also be produced by the Fenton reagent (addition of
H2O2 to Fe2+ salts). Photo-Fenton (Safarzadeh-Amiri et al., 1996) combines Fenton and
UV–VIS light. The photolysis of Fe3+ complexes enables Fe2+ regeneration. Under these
conditions iron can be considered a real catalyst. Both iron and TiO2 catalytic systems are
of special interest because solar light can be used (Malato et al., 2002; Sarria et al., 2005).
References and patents related to the photocatalytic removal of toxic compounds from
129
water published during the last decade can be counted in thousands (Blake, 2001):
however, solar technologies have been most extensively studied and developed for
heterogeneous TiO2 photocatalysis and homogeneous photo-Fenton (Malato and Agüera,
2003). Particularly, investigation upon the photocatalytic oxidation of pesticides in
aqueous media irradiated by near-UV light has been successful and many chemicals were
reported to be degraded efficiently by this method (Malato et al., 1999; Hincapié et al.,
2005; Pichat et al., 2004).
The major limitation of the AOPs is their operational cost. Indeed, for the total oxidation
of hazardous organic compounds or non-biodegradable effluents, financial support
remains relatively high compared to those of a biological treatment (Muñoz et al., 2005).
Therefore it is potentially attractive to exploit these processes as a pre-treatment step to
enhance the biodegradability of wastewater containing recalcitrant or inhibitory
compounds (Scott and Ollis, 1997). Nevertheless this procedure is only justified if the
resulting intermediates are easily degradable by microorganisms in a consecutive
biological treatment because incomplete oxidation may result in more toxic intermediates
than the parents (Amorós et al., 2000; Drzewicz et al., 2004). Therefore the toxicity
assessment of the photodegradation products is the critical point in evaluating the
possibility of photocatalysis to be a pre-treatment process.
Toxicity assessment of a chemical using a single species testing reflects the sensitivity of
that test only: it may over or underestimate the potential toxicity for that particular
substance. Accordingly, recent research has focused on the development of
representative, cost-effective, and quantitative test batteries, which can detect different
effects using a variety of endpoints (Manusadžianas et al., 2003; Davoren and Fogarty,
2004; Escher et al., 2005). However such complete battery of bioassays for mixed
chemical compounds is not yet feasible, in particular for industrial effluents. Furthermore
potential synergistic or antagonist effects may occur and can be problematic to study
(Isnard, 1998; Kungolos et al., 1999; Hauser et al., 1997).
Microbial tests, due to several factors such as short exposure time, ease of handling and
reproducibility of results between laboratories, have been widely used in toxicity
screening procedures. Strains of luminescent bacteria and culture media were studied and
130
evaluated: in 1979, Anthony Bulich proposed a specific test for rapid assessment of the
toxicity of aquatic samples using the light emitting bacterium Vibrio fischeri (former
Photobacterium phosphoreum). In 1982, the system was developed commercially under
the trade mark MicrotoxTM (Berckman Instruments Inc.). Now this test is widely accepted
as a standard for this bioassay type (Parvez et al., 2006; Radix et al., 2000) and has
already been used for evaluation of industrial wastewaters (Lin et al., 1994; Hao et al.,
1996) and study of toxicity of pesticides (Kahru et al., 1996; Jones and Huang, 2003).
Shortening time of phototreatment is of major concern for the cost and energy benefits of
total treatment process (coupling of photocatalysis and biotreatment). The Zahn-Wellens
procedure is the most appropriate method for biodegradation assessment and can be used
for testing biocompatibility of partially photodegraded solutions of pesticides. But this
analytical tool is quite time-consuming (28 days). The main objective of this study was to
evaluate Microtox© as an alternative indicator for biodegradation assessment of partially
phototreated waters contaminated with pesticides. In this study six pesticides (alachlor,
atrazine, chlorfenvinphos, diuron, isoproturon and pentachlorophenol) were considered
separately. Experiments on photocatalytic oxidation of these substances were performed
over photo-Fenton and TiO2, and results are presented. Dissolved organic carbon analysis
was used as a measure of mineralization of pesticides to assess the efficiency of
photocatalytic process. Toxicity assessment of the original pesticides and their
photodegradation products were tested with acute Microtox© testing. Regarding the
evolution of the toxicological level of the photodegraded solutions of pesticides in the
course of the photo-treatment, different stages of photo-Fenton process were taken for
testing their biodegradability with the Zahn-Wellens procedure.
5.A.3. Experimental
5.A.3.1. Chemicals
The chemical compounds used in this study consist of: Alachlor (95% technical grade
C14H20ClNO2, Aragonesas Agro S.A.), atrazine (95%, technical grade C8H14ClN5, Ciba-
Geigy), chlorfenvinphos (93.2%, technical grade C12H14Cl3O4P, Aragonesas Agro S.A.),
131
diuron (98.5%, technical grade C9H10Cl2N2O, Aragonesas Agro S.A.), isoproturon (98%,
technical grade C12H18N2O, Aragonesas Agro S.A.) and pentachlorophenol (98%,
analytical grade C6HCl5O, Aldrich Chemical). Analytical standards of all pesticides (for
chromatographic analyses) were purchased from Sigma-Aldrich. Water used in the pilot
plant was obtained from the Plataforma Solar de Almería (PSA) distillation plant
(conductivity<10µS/cm, Cl- = 0.7-0.8 mg/L, organic carbon <0.5 mg/l). Ferrous iron
sulfate (FeSO4.7H2O), hydrogen peroxide (30% w/v) and sulphuric acid for pH
adjustment (around 2.7-2.8) were reagent grade. TiO2 was Degussa P-25.
5.A.3.2. Analytical determinations
Mineralization was followed by measuring the dissolved organic carbon (DOC) by direct
injection of filtered samples into a Shimazu-5050A TOC analyser provided with an
NDIR detector and calibrated with standard solutions of potassium phthalate. Pesticide
concentration was analysed using reverse-phase liquid chromatography (flow 0.5 ml/min)
with UV detector in a HPLC-UV (Agilent Technologies, series 1100) with C-18 column
(LUNA 5 µm, 3 mm x 150 mm, from Phenomenex). Ultra pure distilled-deionized water
obtained from a Milli-Q (Millipore Co.) system and HPLC-grade organic solvents were
used to prepare all the solutions. Cation concentrations were determined with a Dionex
DX-120 ion chromatograph equipped with a Dionex Ionpac CS12A 4 mm x 250 mm
column. Isocratic elution was done with H2SO4 (10 mM) at a flow rate of 1.2 ml min-1.
Anion concentrations were measured with a Dionex DX-600 ion chromatograph using a
Dionex Ionpac AS11-HC 4 mm x 250 mm column. The gradient programme was pre-run
for 5 min with 20 mM NaOH, injection of the sample, 8 min 20mM NaOH and 7 min
with 35mM of NaOH at a flow rate of 1.5 ml min-1. H2O2 concentration was determined
by iodometric titration.
5.A.3.3. Toxicity Measurements
Microtox© acute toxicity testing was performed with Vibrio fischeri using a Model 500
Analyzer (Azur Environment, Workingham, England). Hydrogen peroxide present in the
samples from photo-Fenton experiments was removed prior to toxicity analysis using
132
catalase (2500 U/mg bovine liver; 100 mg l-1) acquired from Fluka Chemie AG (Buchs,
Switzerland) after adjusting the sample pH to 7. Both samples from photo-Fenton and
TiO2 phototreatments were filtrated (0.22 µm filter, Schleicher and Schuell, G-Dassel)
before toxicity testing. Measurement of toxicity was performed within 24 h after
irradiation. Stored samples should be frozen before analysis. Toxicity is expressed as
toxicity units, TU=100/EC50, where EC50 is the concentration which causes 50%
reduction of the bioluminescence (Vibrio fischeri). All chemicals for the bioassays were
obtained from a commercial supplier (Tetra Technique, Veyrier, Switzerland).
5.A.3.4. Biodegradability assessment
The Zahn-Wellens test was carried out according to the EC protocol (Directive
88/302/EEC). The activated sludge obtained from a secondary effluent of the treatment
plant of Almería (Spain) or Morges (Switzerland) was used as inoculum. The fresh
activated sludge was centrifuged at 10 000 RPM during 7 minutes at 20 °C, and washed
once with mineral medium. According to the guidelines of the Zahn-Wellens test, the
ratio between the carbon content of the experimental sample and the dry-weight of the
inoculum was ranged between 1 and 4 (average 3). Aeration and homogenization were
guaranteed. Preliminary experiments were performed to check that neither volatilization
nor adsorption occurred during the testing period of 28 days.
5.A.3.5. Photoreactor set-up
All the experiments were carried out under sunlight, in a pilot plant at the PSA (latitude
37ºN, longitude 2.4ºW). The pilot plant was operated in batch mode and had three
compound parabolic collectors (CPCs), one tank and one recirculation pump (Malato-
Rodrıguez et al., 2004). Schematic of the photoreactor is illustrated in Figure 5.A.1. Each
collector (1.03 m2 each) was mounted on a fixed platform tilted 37º (local latitude). The
water flowed at 20 l min-1 directly from one module to another and finally into the tank,
from which the pump recirculated the fluid back to the CPCs. The total volume (VT) of
35 l was separated into two parts: 22 l (glass-transparent tubes in the CPC) of total
irradiated volume (Vi) and the dead reactor volume (tank + high density polyethylene
133
tubes), which was not illuminated, as recently described in detail elsewhere (Kositzi et
al., 2004). Solar ultraviolet radiation (UV) was measured by a global UV radiometer
(KIPP&ZONEN, model CUV 3), mounted on a platform tilted 37º (the same as the
CPCs), which provides data in terms of incident WUV m-2. In this way, the energy
reaching any surface in the same position with regard to the sun could be measured.
Based on Eq 5.A.1, the actual time for the irradiation was calculated for the experiment.
1nnnT
in1n30W,n30W, ttt;
V
V
30
UVttt −− −=+= Eq 5.A.1
where tn is the experimental time for each sample, UV the average solar ultraviolet
radiation measured during ∆tn, and t30W the “normalized illumination time”. In this case,
time refers to a constant solar UV power of 30 W m-2 (typical solar UV power on a
perfectly sunny day around noon). This calculation allows the comparison with other
photocatalytic experiments. As the system was outdoors and was not thermally
controlled, the temperature inside the reactor was continuously recorded by a PT-100.
Figure 5.A.1. Flow diagram of photoreactor
134
The characteristics (mainly water solubility) of the six pesticides tested, made the use of
“simulated wastewater” preparation procedures necessary.
Solutions of alachlor, chlorfenvinphos and isoproturon were tested at 50 mg l-1. Since
atrazine, diuron and pentachlorophenol are soluble in water at less than 50 mg l-1 at
ambient temperature, these compounds were tested at the maximum of their
hydrosolubility. These solutions consisted of the filtrated supernatant of a saturated
solution of the target chemical. Exact concentration was measured by HPLC.
During the photocatalytic experiments, DOC analyses were used to monitor the
degradation and mineralization of each pesticide tested in this work. In some cases,
concentration of the inorganic species produced (mainly chloride, nitrate and ammonia)
and direct pesticide concentrations (provided by HPLC) were used for mass balance
calculation.
Concerning the photo-Fenton set of experiments, several prior tests (Hincapié et al.,
2005) were performed with very different Fe concentrations (2 and 55 mg l-1) to find out
the best conditions for comparison with TiO2 and for evaluating toxicity and
biodegradability with both photo-treatments. It has been demonstrated that 2 mg l-1 of Fe
were enough for mineralizing the pesticides under illumination.
5.A.3.6. Data analyses
Validation of toxicological bioassays as well as biodegradability assessments was
ensured since all the quality control data were considered acceptable according to the
official guidelines (OECD, U.S. Environmental Protection Agency (EPA) Office of
Pollution Prevention and Toxics (OPPT), and the European Commission) and other
established criteria (e.g., response to the negative controls, use of reference substances:
phenol for Microtox©, diethylene glycol for Zahn-Wellens testing (Lapertot and Pulgarin,
2006)).
Photocatalytic and biodegradability experiments were at least duplicated and all samples
were analyzed in triplicate.
135
Toxicity data were computed and EC50 values were calculated according to the gamma
method, using linear regression analysis of transformed DOC concentrations as natural
logarithm data versus percentage inhibition. All correlation coefficients were > 0.90.
Statistical analysis of all experimental results was carried out using analysis of variance
(ANOVA). For all laboratory experiments, α was set at 0.05.
Concerning toxicity data, outliers were first detected using Student’s t test. The 5-min
and 15-min EC50 values were compared using one-way ANOVA with Dunett’s multiple
comparison procedure. Control and samples were also compared. Data were tested for
normality and homogeneity of variances. In the case of a failed normality or equal
variance test (e.g. between 5- and 15-min EC50 values or for chlorfenvinphos), data were
rank transformed and tested again. If data were still not normal (it only concerned 5- and
15-min EC50 values), they were compared using a non-parametric Kruskal-Wallis test.
Following the Kruskal-Wallis test, a non-parametric analog of the Dunnett’s procedure
for multiple comparisons was used (Zar, 1996). Finally, each endpoint corresponds to an
averaged value of 4 to 12 validated data.
Such a statistical analysis was performed to strongly confirm the indicated endpoints,
with a significant level of probability (P< 0.05).
5.A.4. Results
5.A.4.1. Photodegradation
The results obtained with the TiO2 and photo-Fenton treatments for each pesticide are
shown in Figures 5.A.2 and 5.A.3, respectively.
Concerning photo-Fenton (Figure 5.A.3), all the experiments were done in the same way.
Once the pesticide dissolved in water had been added to the pilot plant, it was
homogenized in the system.
Then a sample was taken (initial concentration, t30W = -0.75 h) and concentrated sulfuric
acid was added for pH adjustment, so that volume changes could be ignored.
After 15 min a second sample was taken (t30W = -0.5 h) to check if the pH was around 2.7.
136
After that, iron salt (FeSO4.7H2O) was added and mixed for 15 min (t30W = -0.25 h).
Finally, the necessary amount of hydrogen peroxide was added, and after 15 min a
sample was taken to evaluate the dark Fenton reaction (t30W = 0).
At that moment collectors were uncovered and the photo-Fenton experiments began
(“illumination”, Figure 5.A.3a).
In the case of pentachlorophenol, acid was not added in order to avoid precipitation, but
pH evolved very quickly to acidic pH (due to release of chloride) and photo-Fenton was
effective (no Fe precipitation was detected).
In the case of TiO2 tests the initial procedure was shorter.
All the experiments were started recirculating the compound(s) solution, with the
collectors covered, for about 15 min until a constant concentration was achieved
throughout the system.
After sampling, the catalyst, titanium dioxide, was added (final concentration = 200
mg/L) and the mixture recirculated for more than 15 min.
Just before uncovering the CPCs, to start the photocatalytic process, another sample was
collected to characterise the initial condition of the solution.
137
Figure 5.A.2. Evolution of mineralization (DOC), compound removal (HPLC) and
toxicity (100/EC50) during the TiO2 photocatalytic process of single pesticide solutions
tested for alachlor, atrazine, chlorfenvinphos, diuron, isoproturon and pentachlorophenol.
Atrazine
0
20
40
60
80
100
0 1 2 3 4 5
C/C
0 (%
)
0
1
2
3
To
xici
ty (
TU
)
Chlorfenvinphos
0
20
40
60
80
100
0 1 2 3 4 5
C/C
0 (%
)
0
4
8
12
16
20
To
xici
ty (
TU)
Alachlor
0
20
40
60
80
100
0 1 2 3 4 5
t30W (h)
C/C
0 (%
)
0
2
4
6
8
10
12
To
xici
ty (
TU)
HPLC
TOC
100/EC50
Diuron
0
20
40
60
80
100
0 1 2 3 4 5
C/C
0 (%
)
0
510
1520
2530
35
To
xici
ty (
TU
)
Isoproturon
0
20
40
60
80
100
0 1 2 3 4 5
C/C
0 (%
)
051015202530
To
xici
ty (T
U)
Pentachlorophenol
0
20
40
60
80
100
0 1 2 3 4 5t30W (h)
C/C
0 (%
)
0
100
200
300
400
To
xici
ty (
TU
)
138
Figure 5.A.3. Evolution of mineralization (DOC), compound removal (HPLC) and toxicity (100/EC50) during the photo-Fenton process of single pesticide solutions tested
for alachlor, atrazine, chlorfenvinphos, diuron, isoproturon and pentachlorophenol.
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75
t30W (h)C
/C0
(%)
05
1015
2025
3035
To
xici
ty (
TU
)
HPLCTOC100/EC50Zahn-Wellens
Illumination
Alachlor
Al2Al0
Al1
Atrazine
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75
C/C
0 (%
)
0
5
10
15
20
25
30
35
To
xici
ty (
TU
)
At1
At2
At0
Illumination
Isoproturon
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75
C/C
0 (%
)
0
10
20
30
40
50
To
xici
ty (
TU
)
Isp1
Isp2
Isp3Isp0
Diuron
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75
C/C
0 (%
)
0
5
10
15
20
To
xici
ty (
TU
)
D1 D2D0
Pentachlorophenol
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75t30W (h)
C/C
0 (%
)
0
200
400
600
To
xici
ty (
TU
)
Pcp1
Pcp2Pcp0
Chlorfenvinphos
0
20
40
60
80
100
-0.75 0 0.75 1.5 2.25 3 3.75
C/C
0 (%
)
0
150
300
450
600
750
900
To
xici
ty (
TU
)Cl0
Cl2 Cl1
Negative illumination time (t30W) corresponds to the dark period of experiment. Start of illumination is indicated by the vertical line (t30W=0) and the arrow on top of each graph. Points Alx, Atx, Dx, Clx, Ispx, Pcpx indicate samples which were taken for a further biodegradability assessment (See Table V.A.2).
(a)
(b)
(c)
(d)
(e)
(f)
139
It is important to note that during the dark period of photocatalytic treatments neither
mineralization by Fenton nor degradation by TiO2 was observed for all the tested
compounds.
Comparing the time course of photodegradation of pesticides between Figures 5.A.2 and
5.A.3 brings out that in all cases photo-Fenton was quicker than TiO2.
During both photocatalytic treatments, some intermediate products are generated. The
relative amounts of these by-products correspond to the difference between the curves of
HPLC and TOC, on Figures 5.A.2 and 5.A.3. The molecular composition of the partially
degraded solutions of pesticides was not determined but Figures 5.A.2 and 5.A.3 show
that TiO2 and photo-Fenton generate similar proportions of intermediate by-products in
the course of the photodegradation process.
Complete mineralization could be achieved in all cases, except for atrazine.
Comments concerning atrazine and the overall photoreactivity of test pesticides have
been previously detailed and discussed (Farré et al., 2005; Hincapié et al., 2005; Lapertot
et al., 2006b).
5.A.4.2. Toxicity Measurements
The study aimed to assess if the Microtox© analysis could be a useful tool suitable to
detect when a water pre-treated by a phototreatment could be biodegradable.
For that reason Microtox© bioassay was first tested for its reliability to evaluate the acute
toxicity (to Vibrio fischeri) of each pesticide.
Table 5.A.1 gives the EC50 obtained for pure pesticide diluted in MilliQ water after 5-min
exposure. Statistics and literature data for comparison are also listed in Table 5.A.1.
The cumulative results and variation coefficients < 10% indicate that Microtox© bioassay
provides a suitable evaluation of acute toxicity.
Among all the target compounds, it was found that atrazine, alachlor and diuron are the
less toxic substances, whereas pentachlorophenol is the most toxic. As indicated in Table
5.A.1 (see references), obtained data are concomitant with literature. In particular,
atrazine was already designated to be poorly toxic at environmentally relevant
concentrations (Lawton et al., 2005).
140
Table 5.A.1. Experimental and reference data for acute Microtox© EC50 testing for the
target pesticides.
Experimental Literature Test
Compound EC50 (mg/l) Repetition
Variation Coefficient
(%) Reference
EC50
(mg/l)
Alachlor 105 ± 8(1) 6 9.44 Canna-Mickaelidou and Nicolaou, 1996
135 ± 14 (2)
Atrazine 89 ± 6(1) 6 8.44 Boggaerts et al., 2001
Kross et al., 1992
150
13 (3)
Chlorfenvinphos 18 ± 1(1) 6 4.46 Not Found -
Diuron 86 ± 5(1) 8 8.74 Boggaerts et al., 2001 58
Isoproturon 29 ± 2(1) 6 7.09 Parra et al., 2002 25
Pentachloro-phenol
1 ± 0.1(1) 4 5.59
Boggaerts et al., 2001
Kahru et al., 1996
De Zwart and Slooff, 1983
0.7
1.1
0.9
(1): 95% Confidence Interval (mg/l), (2): EC20, 10 min, (3):EC10, 5 min
As illustrated in Figures 5.A.2 and 5.A.3, toxicity was evaluated regularly in the course
of photocatalytic processes. Frequent samples were collected in order to observe the
impact of the oxidative treatment on DOC content, residual pesticide concentration and
toxicological level. Toxicity evolution was considered as an indicator of a progressive
molecular transformation of the solution. It was already demonstrated that variations of
toxicity were representative of the molecular transformations which result from
photocatalytic reactions (Parra et al., 2002) or from other treatments (Drzewicz et al.,
2004) and even biological processes (Lin et al., 1994). As illustrated in Figures 5.A.2 and
5.A.3, each pesticide exhibited a particular toxicological behavior, which also depends on
the applied photocatalytic process (TiO2, photo-Fenton). For instance, overall toxicity
increased at the first steps of oxidation process. But this trend was alternative, depending
on both the type of compound and the photocatalytic process. The toxicity could be
leveled off or up as treatment was proceeding.
141
An increasing toxicity reveals the production of some intermediate metabolites which are
more toxic than the initial molecule (Lapertot et al., 2006a). Thus, whenever the
phototreated solution was considered sufficiently modified regarding its toxicological
variations, the photodegradation was interrupted and a consecutive biodegradability assay
was carried out.
5.A.4.2. Biodegradability testing
Taking into account the higher efficiency of the photo-Fenton with regard to TiO2
process for degrading the target pesticides, it was decided to evaluate biodegradability for
photo-Fenton only.
For each target pesticide, two or three steps were identified and selected for further
biodegradability assessment with the Zahn-Wellens test (Figure 5.A.3, see dark circles
indicated on toxicity curves).
Table 5.A.2 indicates the results obtained for the biodegradability testing with Zahn-
Wellens. The test was considered positive whenever 70% of carbon elimination was
obtained within 28 days maximum according to the EC protocol (Directive 88/302/EEC).
The intermediate results noted (+/-) in Table 5.A.2 correspond to a carbon elimination of
approximately 60% achieved within 28 days, the 70%-threshold being attained later on.
Concerning the biodegradability assessments performed on the initial solutions of single
pesticides, several prior tests were carried out, using either Zahn-Wellens or ZW-slightly
modified methods (Lapertot and Pulgarin, 2006). All pesticides were shown
biorecalcitrant, except atrazine.
Despite the good biodegradability of atrazine, it has been preferred including it in the
study because it has been demonstrated at both lab (Farré et al., 2005) and pilot plant
scales (Hincapié et al., 2005) that when treated by photo-Fenton, the toxicity (measured
by Vibrio fischeri) of the water containing atrazine (around 20 mg/l) increases during the
process, then decreases again after a certain time.
Toxicity was not yet related to biodegradability, but one of the purposes of this paper was
to explore the relationship between these two parameters. Moreover the EU aims to
142
reduce pollution from discharges of hazardous substances (European Commission, 2002),
in which atrazine is concerned.
In this context, any type of AOP applied should demonstrate the complete removal of the
target substance and the harmlessness of the treated effluent.
Table 5.A.2. Biodegradability assessment performed with the partially phototreated
solutions of single pesticide.
Characteristics phototreated solutions
Test Compound Sample(1) Presence of
initial compound
t30W
(h)
DOC/DOCo
(%)
Zahn-Wellens
Result
Alachlor
Al0
Al1
Al2
Yes
Yes
No
0
0
1.36
100
99
16
-
-
+
Atrazine
At0
At1
At2
Yes
Yes
No
0
0
0.75
100
100
78
+
-
+/-
Chlorfenvinphos
Cl0
Cl1
Cl2
Yes
Yes
Yes
0
0
0
100
100
100
-
-
+
Diuron
D0
D1
D2
Yes
No
No
0
0.37
0.81
100
36
15
-
+/-
+
Isoproturon
Isp0
Isp1
Isp2
Isp3
Yes
Yes
No
No
0
0.50
0.66
1.03
100
85
69
30
-
-
+/-
+
Pentachloro-phenol
Pcp0
Pcp1
Pcp2
Yes
No
No
0
1.0
1.27
100
30
11
-
+
+
(1): Samples are taken at the stages of the photo-Fenton treatment indicated in Figure 5.A.3
143
In this study, the solutions resulting from the final stages of both photo-treatments
(photo-Fenton and TiO2) were supposed biocompatible, because either mineralization
was achieved or at least only minimum organic carbon contents were still remaining,
except for atrazine.
For that reason and also because it is aimed to minimize the duration of the photo-
treatment, no biodegradability assessment was performed at the final stages of the
oxidative processes.
From Table 5.A.2 it can be concluded that the photo-treatment period can be significantly
shortened because of the positive Zahn-Wellens results obtained with the partially
degraded solutions of all pesticides.
5.A.5. Discussion
5.A.5.1. Toxicity assessment
To shorten phototreatment time is of major concern for the cost and energy benefits of
total treatment process (coupling of photocatalysis and biotreatment).
Therefore the first aim of this study was to investigate if Microtox© could be considered
as a suitable global indicator capable of giving information on the evolution of
biocompatibility of the water solution contaminated with organic pollutants during the
phototreatment in order to promote biotreatment.
But due to the complexity of the studied process and the specificity and sensibility of the
Microtox© testing, this approach was considered and discussed with caution.
Besides, a more detailed study of the intermediate metabolites as well as other inorganic
species produced such as chloride, ammonium or nitrate contents in addition with toxicity
analyses of these compounds should be achieved in order to improve the knowledge of
the implicated degradation pathways and molecular interactions.
Some studies have already been devoted to similar topics (Hermann et al., 1999; Amorós
et al, 2000).
However tentative statements are proposed in order to explain some specific toxicity
variations.
144
In fact, due to the chemical complexity of the solutions produced during photocatalytic
processes, it seems more interesting to focus on the deviation trend which occurs between
DOC and HPLC data in the course of the photo-degradation (expressed in % C-gram of
the initial concentration) rather than observe the sole overall evolution of toxicity. Indeed,
the difference between the two curves (HPLC and DOC) pictured in Figures 5.A.2 and
5.A.3 indicates the amount of the intermediate metabolites which are generated during
the photocatalytic process.
Considering the removal of the initial molecule on one hand and the disappearance of
DOC on the other hand, two main categories of behaviors can be outlined.
When the produced intermediates demineralize shortly, toxicity decreases regularly in the
course of the photodegradation. This is the case of isoproturon and pentachlorophenol
(photo-Fenton, Figure 5.A.3).
But when the intermediate metabolite degradation takes a long time (after disappearance
of the original compound), level of toxicity is not predictable, but overall in the end, it
tends to level out (for eg. in Figure 5.A.3: atrazine, alachlor, chlorfenvinphos, and
diuron).
As it is illustrated in Figures 5.A.2 and 5.A.3, Microtox© acute toxicity testing
demonstrate the overall dynamics and efficiency of photo-treatment.
5.A.5.2. Biodegradability Assessment
As indicated in Table 5.A.2, for each target compound, several steps of photo-Fenton
treatment were evaluated for biocompatibility.
First of all, as expected, longer process advancement results in a higher potential for the
consecutive biotreatment.
The obtained results show that biodegradability of some compounds can be achieved
even in the presence or absence of the original molecule. For instance alachlor
biodegradation was not observed until alachlor was completely removed by photo-
Fenton. On the contrary, for atrazine and isoproturon, partial degradation results in
biodegradable solution even in presence of original compound. Actually, as shown in
Table 5.A.2, biodegradability was negative for atrazine and isoproturon in the samples
145
where the parent compound was still present. Except for the initial solution of atrazine as
it is a biodegradable compound. Chlorfenvinphos could be also included here.
The obtained results show that chlorfenvinphos was biodegraded even without
illumination. In fact, 80% of chlorfenvinphos was removed during dark-Fenton. It was
transformed into intermediate by-products with no mineralization (DOC constant).
Therefore biodegradability was achieved due to lower amounts of initial compound. It
confirms the results presented in Chapter 2 in which it was stated that chlorfenvinphos
implies irreversible cellular damages that could be overwhelmed whenever the pesticide
concentration was decreased from 50 down to 10 mg l-1 (Lapertot and Pulgarin, 2006).
In case of alachlor, diuron and pentachlorophenol, the progressive decrease of
DOC/DOC0 during photo-Fenton could be related with an increase of biodegradability
(see Table 5.A.2). Indeed biodegradability has been achieved for alachlor, diuron and
pentachlorophenol when DOC/DOC0=16%, 15% and 11% respectively, which means
that the DOC may decrease more than in the case of isoproturon (DOC/DOC0=30%).
For isoproturon, biotreatability of the contaminated water could be finally achieved after
more than 0.7 h of illumination time (t30W, photo-Fenton, Table 5.A.2).
Concerning atrazine, as illustrated in Figure 5.A.3, the intermediate metabolites produced
during photo-Fenton provoked an increasing toxic response to Vibrio fischeri. Besides
after 0.75 h of irradiation, the obtained solution of atrazine was found to be
biorecalcitrant. Surprisingly, the initial solution of pure atrazine has already been proved
biodegradable (Lapertot and Pulgarin, 2006). More precisely, only 40% of carbon content
was eliminated during the Zahn-Wellens test (results not shown). Attending to the results
of DOC, HPLC, nitrogen and chloride analyses, the intermediates were completely
dechlorinated and only the triazine ring was remaining.
As cyanuric acid is known to be biodegradable, other non-chlorinated atrazine
intermediates should be biorecalcitrant (Kontchou and Gschwind, 1999; Horikoshi and
Hidaka, 2003; Lawton et al., 2005). Among the photogenerated intermediates, ammelide
and ammeline are two well-known metabolites produced during photo-treatment of
atrazine (Konstantinou and Albanis, 2003; Chan and Chu, 2005). A specific analysis of
Microtox© acute toxicity performed on these two pure substances showed that their EC50
146
(5-min) is lower than pure atrazine, i.e. 0.12 ± 0.02 and 0.18 ± 0.01 mg l-1, respectively.
This result is concomitant with the observed increasing toxicity during the photo-
treatment of atrazine (Figures 5.A.2 and 5.A.3). Cyanuric acid, which is the final product
formed during oxidation process after long irradiation times (Pellizzetti et al., 1990)
exhibited an EC50 (5-min) of 148 ± 12 mg l-1. The lower toxicity (in comparison with
atrazine and its degradation products) has also been reported by Hiskia et al. (2001).
5.A.6. Conclusions
In this study, it has been shown that photo-Fenton was more efficient than TiO2 for
pesticide degradation and TOC mineralization.
It has also been demonstrated that photo-Fenton treatment is a very consistent method for
enhancing biodegradability of water contaminated with pesticides, and therefore reducing
the costs of the treatment (hydrogen peroxide consumption, photoreactor size, operation
time, etc.).
Indeed biodegradability increases significantly after short photo-Fenton treatment times
for alachlor, diuron and pentachlorophenol.
Uncertain results were obtained with atrazine and isoproturon, which also indicates the
need to perform previous tests before coupling photo-degradation and biotreatment.
The obtained results demonstrate the Zahn-Wellens test as a useful tool for determining
the optimal DOC range in which wastewater becomes biodegradable.
Overall Microtox© acute toxicity testing was shown to represent dynamics and efficiency
of photo-treatment. Hence it was observed that during phototreatment the relative
increase of toxicity could be accompanied by a concomitant increase of biodegradability.
Thus even if Microtox© can not provide a reliable biodegradability assessment by itself, it
can help for selecting samples to be tested by the Zahn-Wellens biodegradability
assessment method.
147
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153
CHAPTER 5.
Integrated Photocatalytic-Biological Process to Treat Pesticides
PART B.
PHOTO-FENTON AND BIOLOGICAL INTEGRATED PROCESS FOR DEGRADATION
OF A MIXTURE OF PESTICIDES
5.B.1. Abstract
A solution of mixed pesticides (alachlor, atrazine, chlorfenvinphos, diuron, isoproturon)
was considered for degradation employing photo-Fenton as a preliminary step before
biotreatment. Photo-Fenton degradation is an important sub-process in the integrated
photobiological process for removal of biorecalcitrant chemicals. Shortening the photo-
Fenton treatment time has a critical impact on the economical feasibility of the integrated
process. In this study, photo-Fenton was proved to enhance biotraitability of a mixture of
biorecalcitrant substances (pesticides). During the photocatalytic process, dissolved
organic carbon (DOC) measurements, liquid and ionic chromatography analyses, acute
toxicity evaluation using MicrotoxTM and the Zahn-Wellens biodegradability testing were
provided to characterize the photo-treated solutions. In order to find out the best moment
for coupling of photocatalytic and biological processes, different times of photo-Fenton
pretreatment were tested for biotreatment. The partially photo-treated solutions were fed
to 8 parallel continuously operated packed-bed bioreactors during 28 days. Considering
the coupled system, it was shown that the pre-treated solutions obtained with only 1 g l-1
Submitted to Journal of Photochemistry and Photobiology, A: Chemistry
154
H2O2 (irradiation time, t30W= 0.6 h) exhibited higher than 80% of DOC degradation. In
this case, the packed-bed bioreactors contributed to more than 50% of the total carbon
conversion of the pesticide solution.
5.B.2. Introduction
Agrochemicals such as insecticides, herbicides and fungicides are classified among the
most dangerous toxicants. Due to their high toxicity, biological treatment of agro-
industrial effluents is often perturbed and sometimes blocked. It is more important to test
the mixture of these substances rather than the individual chemical species because of the
possible presence of combine synergistic and/or antagonistic effects.
In this study, five pesticides were chosen including alachlor, atrazine, chlorfenvinphos,
diuron and isoproturon. These compounds are considered Priority Substances by the EU
(European Commission, 2001) not only from their intrinsic toxicity but also because of
their extremely easy transport in the environment that represents a risk to both surface
and groundwater.
Several oxidative degradation procedures (AOPs, Advanced Oxidation Processes) have
been developed in the field of the chemical treatment of water for the elimination of
pesticides: TiO2/UV, Fenton reagent, O3, O3/UV and O3/H2O2. Of these, photo-Fenton
process was demonstrated particularly efficient for mineralization of pesticides (Farré et
al., 2005 ; Hincapié et al., 2005 ; Malato et al., 1999 ; Pichat et al., 2004). The Fenton
method requires H2O2, Fe2+ salts and acidic pH. Under these conditions highly reactive
and unselective oxidants, OH radicals, are produced. With UV-VIS light, the formed Fe3+
complexes are photolysed, which enables regeneration of Fe2+ (catalyst): the process is
called photo-Fenton.
Photo-Fenton and other AOPs for wastewater treatment have been extensively studied
(Blake, 2001), but their use remains limited due to high operational costs, generation of
UV radiations by lamps, H2O2 consumption in case of photo-Fenton (Pulgarín et al.,
1999 ; Bressan et al., 2004 ; Da Hora Machado et al., 2005). Hence the combination of
photo-Fenton as a preliminary treatment, followed by a biotreatment was considered to
be a feasible process for pesticide removal. Previous papers related the feasibility of the
155
coupling, using the Zahn-Wellens procedure to assess the biocompatibility of the photo-
treated solutions of individual pesticides (Lapertot et al., 2006 ; Sarria et al., 2003a ;
Sarria et al., 2003b). In Part A of Chapter 5, the MicrotoxTM acute toxicity measurement
was considered as a predictive tool to facilitate the biocoupling. In this study, the
biotraitability of the photodegraded solutions was more precisely investigated using
packed-bed bioreactors operated during 28 days.
This paper reports:
(1) the results of the preliminary experiments that demonstrated the biological
incompatibility of the solution of mixed pesticides without any photo-Fenton
pretreatment;
(2) the photo-Fenton degradation of the mixed pesticides. The chemical and
biological characteristics of the photo-treated solutions were evaluated in order to select
different stages of the photocatalytic reaction to be tested for the subsequent
biotreatment;
(3) the results of the biotreatment, which was carried out using several bioreactors
operated in parallel and continuously fed by various pre-treated solutions of mixed
pesticides,
(4) the overall carbon degradation efficiency of the tested photo-Fenton / bioreactor
coupled systems.
5.B.3. Materials and methods
5.B.3.1. Chemicals
The chemical compounds used in this study consist of: alachlor (95% technical grade
C14H20ClNO2, Aragonesas Agro S.A.), atrazine (95%, technical grade C8H14ClN5, Ciba-
Geigy), chlorfenvinphos (93.2%, technical grade C12H14Cl3O4P, Aragonesas Agro S.A.),
diuron (98.5%, technical grade C9H10Cl2N2O, Aragonesas Agro S.A.) and isoproturon
(98%, technical grade C12H18N2O, Aragonesas Agro S.A.). Analytical standards of all
pesticides (for chromatographic analyses) were purchased from Sigma-Aldrich.
156
Demineralised water was used to prepare the mixture of pesticides. Ferrous iron sulfate
(FeSO4.7H2O), hydrogen peroxide (30% w/v) and sulphuric acid used for pH adjustment
(around 2.7-2.8) were reagent grade.
5.B.3.2. Analytical determinations
Mineralization was followed by measuring the dissolved organic carbon (DOC) by direct
injection of filtered samples into a Shimazu-5050A TOC analyser provided with an
NDIR detector and calibrated with standard solutions of potassium phthalate. Pesticide
concentration was analysed using reverse-phase liquid chromatography (flow 1 ml/min)
with UV detector in a HPLC-UV (Varian 9012, 9100, 9065) with an ODS-2 column
(Waters 4.6x250 mm, from Phenomenex) and a guard column (Waters 4.6x10 mm):
Alachlor (H2O/Acetonitrile 40/60, 224nm), Atrazine, Isoproturon, Chlorfenvinphos and
Diuron (Acetic acid(1%)/Acetonitrile 80/20-40/60 (0-30 min) and 40/60 (30-35 min), 249
nm). Ultra pure distilled-deionised water obtained from a Milli-Q (Millipore Co.) system
and HPLC-grade organic solvents were used to prepare all the solutions. Cation
concentrations were determined with a Dionex DX-120 ion chromatograph equipped with
a Dionex Ionpac CS12A 4 mm x 250 mm column. Isocratic elution was done with H2SO4
(10 mM) at a flow rate of 1.2 ml min-1. Anion concentrations were measured with a
Dionex DX-600 ion chromatograph using a Dionex Ionpac AS11-HC 4 mm x 250 mm
column. The gradient programme was pre-run 5 min with 20 mM NaOH, injection, 8 min
20mM NaOH and 7 min NaOH 35mM, flow rate 1.5 ml min-1. H2O2 concentration was
determined by iodometric titration.
5.B.3.3. Toxicity measurements
MicrotoxTM acute toxicity testing was performed with Vibrio fischeri using a Model 500
Analyzer (Azur Environment, Workingham, England). Hydrogen peroxide present in the
samples from photo-Fenton experiments was removed prior to toxicity analysis using
catalase (2500 U/mg bovine liver; 100 mg l-1) acquired from Fluka Chemie AG (Buchs,
Switzerland) after adjusting the sample pH to 7. Samples from photo-Fenton treatment
were filtrated (0.22 µm filter, Schleicher and Schuell, G-Dassel) before toxicity testing.
Measurement of toxicity was performed within 24 h after irradiation. Stored samples
157
should be frozen before analysis. Toxicity is expressed as toxicity units, TU=100/EC50,
where EC50 is the concentration which causes 50% reduction of the bioluminescence
(Vibrio fischeri). All chemicals for the bioassays were obtained from a commercial
supplier (Tetra Technique, Veyrier, Switzerland).
5.B.3.4. Biodegradability assessment
The Zahn-Wellens test was carried out according to the EC protocol (Directive
88/302/EEC). The activated sludge obtained from a secondary effluent of the treatment
plant of Morges (Switzerland) was used as inoculum. The fresh activated sludge was
centrifuged at 10 000 RPM during 7 minutes at 20 °C, and washed once with mineral
medium. According to the guidelines of the Zahn-Wellens test, the ratio between the
carbon content of the experimental sample and the dry-weight of the inoculum was
ranged between 1 and 4 (average 3.2). Aeration and homogenization were guaranteed.
Preliminary experiments were performed to check that neither volatilization nor
adsorption occurred during the testing period of 28 days.
5.B.3.5. Experimental set-up
5.B.3.5.1. Photo-Fenton experiments
The photocatalytic experiments were performed using 0.5 l Pyrex flask with a cut-off at
λ=290 nm placed into Hanau Suntest Simulator.
The radiation source employed was a xenon lamp and the total radiant flux (80 mW cm-2)
was measured with a YSI Corporation powermeter. The lamp had a λ distribution with
about 0.5% between 300 and 400 nm. The profile of the photons emitted between 400
and 800 nm followed the solar spectrum. Based on Eq.5.B.1, the actual time for the
irradiation was calculated for the experiment.
1nnnn1n30W,n30W, ttt;30
UVttt −− −=+= Eq 5.B.1
where tn is the experimental time for each sample, UV the average solar ultraviolet
radiation emitted during ∆tn, and t30W the “normalized illumination time”. Thus time
158
refers to a constant solar UV power of 30 W m-2 (typical solar UV power on a perfectly
sunny day around noon). This calculation allows the comparison with other
photocatalytic experiments.
The aqueous suspensions were magnetically stirred throughout irradiation, opened to air.
Extreme care was taken to ensure uniform experimental conditions during the partial
degradations, since several experiments were necessary to be run in order to continuously
feed the bioreactors.
5.B.3.5.2. Biotreatment experiments
As illustrated in Figure 5.B.1, 8 aerated packed-bed bioreactors were operated in parallel.
Each bioreactor consisted of a 0.24 l-glass column containing ca. 0.15 l packing
expanded clay colonized by activated sludge from the wastewater treatment plant of
Morges (Switzerland). Liquid in the columns was circulated with peristaltic pumps.
The effluent of the photocatalytic stage was circulated through the bottom of the column,
which operated as an up-flow reactor. In order to homogenize the bacterial population
throughout each column and between all of them (in fact, 8 biofilters were run
simultaneously), wastewater from the primary decantor was recirculated inside the
biofilter from upside down during a few days.
The pH was controlled and adjusted at 7 using 1M H2SO4 or 2 M NaOH solutions.
The required nutrients for bacterial activity were provided after photo-treatment with a
concentrated mineral medium, so that volume remained unchanged.
The aeration was about 0.03 l h-1 and the O2 concentration was checked by means of a
dissolved oxygen (DO) probe (Ingold AG, Urdorf, Switzerland) on top of the column. All
the glass vessels were protected from illumination with aluminum sheets. Agitation was
maintained before and after biotreatment using magnetic stirring. Temperature was 25°
C.
159
Figure 5.B.1. Experimental setup used for investigation of the photocatalytic-
biological coupling treatment of the mix of pesticides.
5.B.3.6. Data analyses
Validation of toxicological bioassays as well as biodegradability assessments was
ensured since all the quality control data were considered acceptable according to the
official guidelines (OECD, U.S. Environmental Protection Agency (EPA) Office of
Pollution Prevention and Toxics (OPPT), and the European Commission) and other
established criteria (e.g., response to the negative controls, use of reference substances:
phenol for MicrotoxTM, diethylene glycol for Zahn-Wellens testing).
Photocatalytic and biodegradability experiments were at least duplicated and all samples
were analyzed in triplicate. Toxicity data were computed and EC50 values were calculated
Outlet liquid
Inlet pretreated
liquid
Magnetic stirring
Glass porous plate
Bioreactors : packing expanded clay in
glass columns
Teflon piping
Silicone piping
Inlet air
160
according to the gamma method, using linear regression analysis of transformed DOC
concentrations as natural logarithm data versus percentage inhibition. All correlation
coefficients were > 0.90.
Statistical analysis of all experimental results was carried out using analysis of variance
(ANOVA). For all laboratory experiments, α was set at 0.05.
5.B.4. Results and discussion
5.B.4.1. Biocompatibility of the mixed pesticide solution
Before considering any photocatalytic treatment of the mixture of pesticides, a solution of
the mixed compounds (alachlor, atrazine, chlorfenvinphos, diuron, isoproturon) was first
tested for biodegradability using the Zahn-Wellens procedure (see Chapter 2). This test
was carried out under similar conditions that of a wastewater treatment plant using
activated sludge. A parallel control experiment using diethylene glycol (0.5 g/l) was
carried out to test if the sludge was active. The diethylene glycol was in fact degraded as
much as 96% in 6 days.
Under the tested conditions, the mixture of pesticides was proved biorecalcitrant. The
concentrations of DOC and of each pesticide (HPLC) in the mix remained unchanged,
even after 28 days. This observation also indicates that bacteria cannot adapt to degrade
these chemicals. This result is concomitant with previous experiments performed with the
single compounds and using the same biodegradability testing procedure (Chapter 2). In
that study, no biodegradation has been observed for each compound tested alone, except
for atrazine (see Table 5.B.1).
A supplementary test to measure the biodegradation of the mix of pesticides was carried
out in batch mode in the packed-bed bioreactors shown in Figure 5.B.1.
Even under theoretically favorable conditions, such as the presence of co-substrates and
adapted bacteria, as well as a strict control of pH, temperature and aeration, the test
confirmed that the solution of mixed pesticides is non-biodegradable in the tested
conditions.
161
Three techniques were used to follow this test:
(a) respirometric measurements with an O2 probe in both the inlet and outlet of the
biofilters,
(b) determination of single pesticide concentrations by HPLC and
(c) measurements of DOC as a function of time.
Table 5.B.1. Physico-chemical and biological properties of the pesticides.
(NB: Non-Biodegradable, B: Biodegradable)
5.B.4.2. Chemical and biological characteristics of the photo-treated pesticide solution
Photo-Fenton degradation is an important sub-process in the integrated photocatalytic-
biological coupled process for removal of biorecalcitrant chemicals.
Shortening the photo-Fenton treatment time has a critical impact on the economical
feasibility of the integrated process.
For this purpose, it is very important to gather information concerning both chemical and
biological characteristics of the solution during the photocatalytic pre-treatment.
DOC analyses, concentrations of the initial compounds and of other inorganic species
produced during the photo-Fenton process were monitored in the course of the
photocatalytic experiments.
Compound Molecular
formula
Solubility
in water at
25°C
(mg/l)
EC50 ± 95%
Confidence
Interval (mg/l)
Biodegradability
(Zahn-Wellens
test)
Initial
concentration
(mg/l)
Alachlor C14H20ClNO2 239 105 ± 8 NB 30
Atrazine C8H14ClN5 34 89 ± 6 B 30
Chlorfenvinphos C12H14Cl3O4P 125 18 ± 1 NB 30
Diuron C9H10Cl2N2O 41 86 ± 5 NB 30
Isoproturon C12H18N2O 65 29 ± 2 NB 30
162
Acute toxicity was also measured using the MicrotoxTM system and biodegradability was
assessed according to the Zahn-Wellens method.
Decreasing DOC concentrations are representative of the oxidative reactions that occur in
the experimental solution during the photocatalytic treatment.
As shown in Figure 5.A.2, evolution of global mineralization can be well illustrated by
both H2O2 consumption and the illumination time. Therefore the progress of the photo-
Fenton process could be equally defined by one of the two parameters. In fact, the
concentration of H2O2 was precisely followed in order to avoid any residual amount of
H2O2 in the photo-treated solution, because H2O2 may damage the bacterial cells and thus
limit the consecutive biotreatment. No remaining H2O2 was also required for toxicity
analyses.
Figure 5.B.2. DOC evolution of the solution of mixed pesticides as a function of the
irradiation time (t30W) and the consumed amounts of H2O2 during photo-Fenton.
Error bars represent 95% confidence intervals.
Figure 5.B.2 shows that the degradation ratio tends to stabilize at H2O2 concentrations
above 2 g l-1. Beyond this threshold, higher amounts of H2O2 as well as a higher
irradiation time are necessary to extend degradation.
In fact, the mineralization of the mixed pesticide solution was difficult to achieve.
Long irradiation period and its consequent high time and energy consumption, as well as
the high H2O2 amounts result in a rather expensive treatment. Electricity represents about
y = -29.73x + 95.23
R2 = 0.91
y = -31.47x + 85.21
R2 = 0.910
25
50
75
100
0 1 2 3 4[H2O2] / t30W
C/C
0 (%
)
H2O2 (g/l)
t-30W (h)
163
60% of the total operational cost when photoprocesses are driven by electrical generation
of photons instead of direct solar irradiation.
This underlines the aim of the study to shorten the photocatalytic treatment time in order
to promote the biodegradation.
In this study, the illumination time (t30W, described in section 5.B.3.5.2.) was chosen to
characterize the advancement of the photo-Fenton process, because it is a useful
parameter in practical applications.
Figure 5.B.3 represents the chemical degradation of the mix of pesticides as a function of
the illumination time course (t30W). Evolution of the overall MicrotoxTM acute toxicity is
also pictured, toxicity being expressed in TU (=100/EC50, Lankford and Eckenfelder,
1990).
Figure 5.B.3. Evolution of toxicity (expressed as 100/EC50) and DOC of the solution of
the mixed pesticides as a function of the irradiation time (t30W) during photo-Fenton.
Data correspond to the average value of at least six measurements.
A sharp decrease in toxicity was observed at the beginning of the photo-Fenton process,
followed by a stable toxic level during all the rest of the photo-treatment.
This indicates that the intermediates formed during the photo-assisted pre-treatment are
less toxic than the initial compounds.
This is different from the behavior observed with the photo-Fenton degradation of the
single pesticides (Chapter 5, Part A). Indeed, even if many fluctuations have been
0
20
40
60
80
100
-0.75 0.00 0.75 1.50 2.25 3.00t30W (h)
C/C
0 (%
)
01
23
45
67
To
xici
ty (
TU
) DOC
100/EC50
Illumination
164
mentioned, mainly depending on the intermediate by-products generated by each tested
compound, it has been shown that toxicity of single pesticide solutions tends to increase
during the first stages of the photo-treatment and then decreases.
So it seems that whenever the substances are mixed, toxicity presents a global behavior
which tends to decrease as soon as the photo-Fenton begins. It has already been stated
that toxicity is not an additive data (Kahru, et al., 1996).
Moreover, as illustrated in Table 5.B.2, comparison between the irradiation times
necessary to degrade the single and the mixed compounds reveals that the photo-Fenton
process is more efficient for treating the mixing solution, even if incomplete
mineralization was achieved.
Table 5.B.2. Comparison of the illumination time (t30W) necessary to remove 100% of the
parent molecule of pesticide (HPLC data) and 80% of DOC during the photo-Fenton
treatment of both single and mixed pesticides.
The evolution of the 5 initial compounds was followed during the pre-treatment process
by HPLC measurements.
At 0.15 h of photo-treatment, when about 20% of DOC removal was observed, mainly
every parent pesticide was eliminated (only 3% atrazine and 4% chlorfenvinphos were
Illumination time,t30W (in h)
100 % removal of initial compound 80 % DOC removal Compound
Single In mixture Single In mixture
Alachlor 0.28 ± 0.01 0.05 ± 0.04 1.20 ± 0.02
Atrazine 0.94 ± 0.05 0.85 ± 0.03 >3.50
Chlorfenvinphos 0.25 ± 0.01 0.05 ± 0.01 1.13 ± 0.03
Diuron 0.37 ± 0.02 0.25 ± 0.01 0.68 ± 0.02
Isoproturon 0.66 ± 0.03 0.50 ± 0.02 1.28 ± 0.30
2.50 ± 0.02
165
left). Thus complete disappearance and total dechlorination (23 mg l-1 chloride was
expected for the tested concentrations) of all pesticides were attained very easily.
The elimination of the initial bio-recalcitrant compound was required in order to test the
biocompatibility of the photo-treated solution (Parra et al., 2002).
The slow decrease of DOC compared with the pesticide evolution indicated an
accumulation of intermediate products.
The nitrogen balance was also checked during the photo-Fenton process, because the
release of ammonia and nitrate was indicator of the by-product degradation.
The slow mineralization rate observed at the end of the photo-treatment was concomitant
with the incomplete release of nitrogen as NH4+ or NO3
-.
From previous experiments (Chapter 5, Part A), it has been concluded that the most
resistant intermediates were due to the phenylurea pesticides (diuron and isoproturon)
and atrazine. In particular, the stable triazine ring and the formation of urea could justify
the remaining 10% DOC measured at the end of the photo-Fenton process.
In the case of the mixed pesticide solution, toxicity and biodegradability were found to be
related.
Indeed, as illustrated in Figure 5.B.3, before irradiation (i.e. during the Fenton process),
the solution of the mixed pesticides was toxic up to 6 TU and was also biorecalcitrant
according to the test of Zahn-Wellens.
On the contrary, after illumination (i.e. during the photo-Fenton process) the partially
treated solutions exhibited a lower toxicity, approximately 0.5 TU (see Figure 5.B.3), and
were biodegradable according to the Zahn-Wellens test.
These results are in accordance with previous experiments which demonstrated that
biocompatibility was achieved after complete disappearance of the initial biorecalcitrant
substance (Lapertot et al., 2006).
Therefore it is interesting to test the biotreatability of the solutions partially degraded by
the photo-Fenton process.
166
5.B.4.3. Photochemical-biological coupled flow treatment
The above-mentioned results presenting the concomitant decrease of toxicity and increase
of biodegradability of the treated solution of mixed pesticides suggest the photo-Fenton
as a promising pre-treatment process.
Therefore, a photochemical-biological coupled flow reactor can be considered for the
complete mineralization of the solution of the mixed pesticides.
The aim of the study was also to determine the best moment for coupling, in order to
limit the cost of the photocatalytic treatment.
For that purpose, successive steps of pre-treatment were selected during the photo-Fenton
process and tested for biotreatment in several biological fixed bed reactors.
The bioreactors (see Figure 5.B.1) were operated in continuous mode, whereas the photo-
chemical treatment was operated in batch mode.
Table 5.B.3 gives the characteristics of different pre-treated solutions which were tested
for biocoupling.
Table 5.B.3. Photocatalytic characteristics and biodegradability assessment of the pre-
treated solutions tested with the photo-Fenton / Bioreactor coupled system.
(NB: Non-Biodegradable, PB: Partially-Biodegradable, ND: Not-Determined)
Sample t30W (h) H2O2 (g l-1) C/C0 (%) Biodegradability (Zahn-Wellens)
A 0 0 96 NB
B 0.11 0.50 85 PB
C 0.25 0.75 76 PB
D 0.60 1.00 60 PB
E 0.80 1.23 48 ND
F 1.20 2.00 42 ND
G 1.76 2.50 35 ND
H 2.06 3.75 26 PB
167
Each partially-treated solution was tested in triplicate: 3 columns were operated in
parallel and were compared to the control bioreactors.
Control bioreactors consisted of a bioreactor fed with only the mineral medium and a
bioreactor operated with the initial solution of no photo-treated mixed pesticides.
A preliminary test using diethylene glycol was performed in each bioreactor in order to
check both the activity of the bacteria and the correct operating mode of each column.
The pre-treatment of the solution was made in photocatalytic reactors using the photo-
Fenton process.
After photo-treatment, the solutions were neutralized (pH around 6.5-7), which also
provoked the precipitation of the iron. Solutions were centrifuged and filtrated before
biotreatement.
Sequential batches of the photo-treatment process were carried out so that the input flow
rate into the biological reactors was 0.12 l d-1. This flow rate was maintained during at
least 28 days. A relatively steady-state condition for DOC removal was observed in the
columns after 48 hours.
Various experiments were performed in order to compare the efficiency of the coupled
photo-catalytic and biological treatments of the mixture of pesticides based on the
advancement of photodegradation.
Figure 5.B.4 illustrates the experimental results obtained with the coupled system. Data
correspond to the average values obtained after 28 days of biotreatment in the columns.
Figure 5.B.4 shows the DOC removal efficiency which was attributed to both photo-
Fenton and biotreatment, as a function of the photoreaction time presented in Table
5.B.3.
168
Figure 5.B.4. Relative contribution of photocatalysis and biotreatment for the degradation
of the solution of the mixed pesticides. Characteristics of the samples A-H are listed in
Table 5.B.2. Error bars represent 95% confidence intervals.
A maximum degradation efficiency of 90% was obtained with the coupling system. It
was achieved with pre-treated solutions E, F, G and H.
Thus beyond an irradiation time of 0.80 h (1.25 g l-1 H2O2, sample E), intensifying the
photo-Fenton reaction is useless to enhance the global degradation ratio. So it means that
beyond this threshold, the more intensive the photo-treatment, the less important the
biological contribution to achieve the global degradation.
In fact, more than 80% of DOC were degraded in the coupled system with pre-treated
solutions using only 1 g l-1 H2O2 (t30W=0.6 h, sample D). This global degradation degree
is already an interesting result for an industrial application of the integrated process. In
this case, the biological treatment contributes to more than 50% of the total carbon
conversion of the pesticide solution.
It should be noticed that using a recirculation loop in biological system can have
important effect on the total removal capacity and intensification of the integrated
process. Since the toxicity of the pesticide solution becomes low after partial photo-
Fenton process, the adaptation of bacteria is expected to enhance the biotreatment.
0
25
50
75
100
A B C D E F G H
Deg
rad
atio
n r
atio
(%
)
Biotreatment
Photo-Fenton
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5.B.5. Conclusion
The photocatalytic treatment of the solution of the mixed pesticides was successfully
achieved with the photo-Fenton process. In this study, photo-Fenton degradation was
considered as a sub-process in the integrated photocatalytic-biological coupled system for
removal of the biorecalcitrant pesticides. In particular, it was aimed to shorten the photo-
treatment time in order to enhance the economical feasibility of the integrated process.
The MicrotoxTM acute toxicity evaluation and the Zahn-Wellens tests for biodegradability
assessment were shown to be two relevant indicators to determine the best moment for
coupling. Biotreatability of the partially photo-degraded solutions was investigated more
accurately using several biological fixed bed reactors which were continuously operated
during 28 days.
The obtained concomitant decrease of toxicity and increase of biodegradability of the
partially photo-treated solution of mixed pesticides confirmed that photo-Fenton is a
promising pre-treatment process capable to enhance the biotreatability of waters
contaminated with biorecalcitrant chemicals (pesticides).
Finally more than 80% of DOC was degraded in the coupled system for the pesticide
solutions which were photo-treated during only 0.6 h (t30W) and needed 1 g l-1 H2O2. This
step of photo-Fenton treatment proved to be the most relevant moment for coupling
because in this case, the biological treatment contributed to more than 50% of the total
carbon conversion of the pesticide solution. Recirculation of the bioreactor effluents is
now prospected in order to improve the global degradation degree of the contaminated
waters.
5.B.6. References
Blake, D.M., 2001. Bibliography of work on the photocatalytic removal of hazardous
compounds from water and air. National Technical Information Service, US Dep. of
Commerce, Springfield, VA, USA.
170
Bressan, M., Liberatorre, L., D’Alessandro, N., Tonucci, L., Belli, C., Ranalli,G., 2004.
Improved combined chemical and biological treatments of olive oil mill wastewaters. J.
Agric. Food Chem. 52, 1228-1233.
Da Hora Machado, A.E., Xavier, T.P., de Souza, D.R., de Miranda, J.A., Mendonsa-
Duarte, E.T.F., Ruggiero, R., de Oliveira, L., Sattler, C., 2005. Solar photo-Fenton
treatment of chip board production waste water. Solar Energy 77, 583-589.
European Commission, 2001. Decision No 2455/2001/EC of the European Parliament
and of the Council of 20 November 2001 establishing the list of priority substances in
the field of water policy and amending Directive 2000/60/EC.
Farré, M.J., Franch, M.I., Malato, S., Ayllón, J. A., Peral, J., Domènech, X., 2005.
Degradation of some biorecalcitrant pesticides by homogeneous and heterogeneous
photocatalytic ozonation. Chemosphere 58, 1127-1133.
Hincapié, M., Maldonado, M.I., Oller, I., Gernjak, W., Sánchez-Pérez, J.A., Ballesteros,
M.M., Malato, S., 2005. Solar photocatalytic degradation and detoxification of EU
priority substances. Catalysis Today 101, 203-210.
Kahru, A., Tomson, T.P., Külm, I., 1996. Study of toxicity of pesticides using
luminescent bacteria Photobacterium phosphoreum. Wat. Sci. Tech. 33 (6), 147-154.
Lapertot, M., Pulgarín, C., 2006. Biodegradability Assessment of Several Priority
Hazardous Substances: Choice, Application and Relevance Regarding Toxicity and
Bacterial Activity. Chemosphere, In press
Lapertot, M., Pulgarin, C., Fernández-Ibáñez, P., Maldonado, M. I., Pérez-Estrada, L.,
Oller, I., Gernjak, W., Malato, S., 2006. Enhancing biodegradability of priority
substances (pesticides) by solar photo-Fenton. Wat. Res. 40 (5), 1086-1094.
Lankford, P., Eckenfelder, W., 1990. Toxicity Reduction in Industrial Effluents. Van
Vostrand Reinhold, New York, USA.
Malato, S., Blanco, J., Richter, C., Milow, B., Maldonado, M.I., 1999. Pre-industrial
experience in solar photocatalytic mineralization of real wastewaters. Application to
pesticide container recycling. Wat. Sci. Tech. 40 (4-5), 123-130.
171
Parra, S., Pulgarín, C., Malato, S., 2002. New integrated photocatalytic-biological flow
system using supported TiO2 and fixed bacteria for the mineralization of isoproturon.
Appl. Catal. B: Environ. 36, 131-144
Pichat, P., Vannier, S., Dussaud, J., Rubis, J.-P., 2004. Field solar photocatalytic
purification of pesticides-containing rinse waters from tractor cisterns used for
grapevine treatment. Solar Energy 77 (5), 533-542.
Pulgarín, C., Invernizzi, M., Parra, S., Sarria, V. Polaina, R., Péringer, P., 1999. Strategy
for the coupling of photochemical and biological flow reactions useful in mineralization
of biocalcitrant industrial pollutants. Catalysis Today 54, 341-352.
Sarria, V., Deront, M., Péringer, P., Pulgarín, C., 2003a. Degradation of biorecalcitrant
dye precursor present in industrial wastewater by a new integrated iron(III)
photoassisted-biological treatment. Appl. Catal. B: Environ. 40, 231-246.
Sarria, V., Kenfack, S., Guillod, O., Pulgarin, C., 2003b. An innovative coupled solar-
biological system at field pilot scale for the treatment of biorecalcitrant pollutants J.
Photochem. Photobiol. A: Chemistry 159, 89–99.
172
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CHAPTER 6.
Conclusion and perspectives of the thesis
During the XXth Century, legislation has been developed to protect the quality of the
limited environmental resources. However, obvious and repetitive ecological catastrophes
reveal the urgency to react efficiently.
Overall more stringent pollution controls have resulted in promotion of research topics
related with pollution removal. For this, up-to-date approaches for industrial processes as
well as for pollution treatments now need to be ecologically sustainable.
In that sense, this thesis was aimed to propose a strategy for treating problematic
pollutants with the most adequate process.
In Chapter 1, an overview of the different treatment processes has been given. Among the
existing processes, biological treatment of contaminated environments and industrial
effluents has been presented as the most attractive method on a cost-benefit extent. The
Advanced Oxidation Processes (AOP), and particularly TiO2 and photo-Fenton
photocatalytic treatments, have also been described as they are really efficient for
removal of recalcitrant substances. But the economical drawback of such processes has
contributed to develop new hybrid technologies combining both AOP and biological
processes for the removal of recalcitrant and/or toxic substances. Therefore in this thesis,
the three above-mentioned processes - biological, photocatalytic and integrated
biological-photocatalytic processes - have been tested for degradation of target
xenobiotics. These chemical compounds were pesticides, phenolic substances and
volatile organic compounds (VOCs). The schematic of the general strategy proposed in
this thesis has been illustrated in this Chapter 1.
174
The first step of the proposed strategy has intended to evaluate biotraitability for 19
chemicals, which were pesticides, pharmaceuticals and volatile organic compounds
(VOCs).
Therefore Chapter 2 was aimed to enhance the procedure of biodegradability assessment,
so that both volatile and hydrophobic substances could be tested. It has been shown that
the most suitable test of biodegradability could be selected and adapted in accordance
with the physicochemical properties of each compound. In particular, three official
protocols - Zahn-Wellens, Manometric Respirometry and Closed-Bottle tests - have been
adapted and performed. They have permitted to evaluate the biotreatability of xenobiotics
in industrial wastewater treatment plant conditions. Acute toxicity (MicrotoxTM),
bacterial viability (Bac-lightTM) and Structure Activity Relationships (SARs) models
have also been used to investigate the influence of each substance on the microbial
population (sludge).
The main contribution of this study has been to adapt biodegradability assessment
methods for the testing of chemical compounds which were considered unsuitable
regarding the existing protocols. The comparison between the screening level
experiments and the several SARs predictive models has also been performed, so that
experimental results have agreed with estimates for 16 xenobiotics. Finally, this study has
permitted to demonstrate that comparison between several estimation models was
beneficial in order to limit overall environmental risks.
The next step of the proposed strategy has depended on the results of the biodegradability
test procedure presented in Chapter 2.
A positive result means that the target compound can be degraded by microorganisms.
Among the 19 xenobiotics which have been tested in Chapter 2, the VOCs have revealed
biodegradable. Therefore their biodegradation has been studied in order to improve the
treatment process.
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In Chapter 3, the Substrate Pulse Batch (SPB) process has been presented as a suitable
technique for cultivation of mixed bacteria capable to degrade chlorinated aromatics and
TEX (toluene, ethylbenzene, xylenes) compounds. It has been shown that batch
cultivation feeded by pulses of substrate (SPB technique) permitted to overwhelm both
toxicity and inhibition of VOCs, so that biomass production could be enhanced. Indeed,
in chapter 3 part A, 240 mg l-1 h-1 of CB/DCB mixture has been eliminated and 50 mg l-1
h-1 of dry weight have been produced using SPB process. In particular, studying substrate
removal dynamics, concentration of metabolic intermediates, chlorides and cellular
integrity has allowed the correct proceeding of bacterial cultivation. The time interval
between each pulse of substrate has been decreased gradually during the CB/DCB
biodegradation process, so that a continuous substrate feeding has finally been attained.
In the continuous feeding technique, both cell productivity and CB/DCB elimination
capacity have been twice the values obtained with the SPB. In chapter 3 part B,
automation of the process has been successfully achieved for removal treatment of TEX
compounds. A monitoring program based on LabVIEW software has been developed in
order to automate the process and to limit the operator contribution. In Part B, the
dynamic behavior of the mixed bacterial population has also been observed and discussed
under different operational conditions. Carbon and nitrogen limitations have been shown
to affect the integrity of the bacterial cells as well as their production of exopolymeric
substances (EPS). Average productivity and yield values have reached 0.45 kgDW m-3 d-1
and 0.59 gDW gC-1, respectively. Finally, obtained data have come up to the industrial
specifications and have indicated the feasibility of controlled SPB technique for TEX
removal.
As for CB/DCB removal, it would be interesting to perform a continuous process for
TEX biodegradation with the automated technology. More specifically, it could be useful
to clarify the dynamics of the bacterial population, in order to avoid the observed
decrease in degradation rates which have systematically occurred during the biological
removal of VOCs. For instance, it is assumable that the composition of the bacterial
consortium changes along the process. Indeed specific strains are certainly more capable
to tolerate either the initial substrates or the intermediate metabolites. Genetic adaptations
176
might certainly be associated with the evolution of the substrate degradation kinetics.
Tools of molecular biology could provide the expected information.
In case of negative results to biodegradability assessment, the biorecalcitrant compound
needs to be treated by other treatment technologies.
TiO2 and photo-Fenton photocatalytic treatments are two efficient and ecologically
friendly advanced oxidation processes (AOPs). They have been considered in Chapters 4
and 5.
Chapter 4 has focused the degradation of halogenated phenolic substances by means of
TiO2 photocatalysis. The influence of the halogen upon the degradation of p-halophenols
in water has been investigated. Phenol has been used as the reference compound.
Compared with its value for phenol, the apparent first-order rate constant of removal, k,
has been shown slightly but significantly higher for p-fluorophenol and p-chlorophenol,
and slightly but significantly lower for p-bromophenol. For p-iodophenol, k has been
about half that of phenol. The relative values have thus confirmed that k was roughly
correlated to the Hammett constant. Conversely, since all compounds have been poorly
adsorbed on TiO2, their various degradation rates have not been related to the difference
of halogen. The organic intermediate products have been detected; they included
hydroquinone (HQ), benzoquinone (BQ) and various halodihydroxybenzenes.
Mechanisms have been tentatively suggested to interpret the differences in the
degradation pathways of the p-halophenols. The MicrotoxTM acute toxicity has also been
investigated during the TiO2 photocatalytic removal of p-halophenols. Interpretations
have been tentatively suggested to link the observed variations in toxicity and the
intermediate metabolites produced according to the halogen. Finally, even though the
total quantity of these aromatic intermediate products has been found to be similar
regardless of the halogen, at equal removal percentage of p-halophenol the variations in
toxicity has markedly differed for solutions that initially contained p-iodophenol.
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This study has illustrated how the molecular structure of a chemical could influence its
reactivity. However, since the molecular composition becomes more and more complex
along the degradation process, further investigations are required in order to provide
more relevant assessments about the overall variations in toxicity and the influence of the
halogen upon this parameter.
In Chapter 5, the combination of photocatalytic and biological systems has been explored
for treatment of water contaminated with bioreacalcitrant pesticides. For the development
of this coupled process, several points have been considered:
o the biotraitability of the initial pollutant,
o the chemical and biological characteristics of the photo-treated solutions,
o the definition of the optimal pretreatment time,
o the efficiency of the coupled process.
Shortening phototreatment time has been of major concern for the cost and energy
benefits of the xenobiotics degradation performed by photocatalytic processes. Using
photo-Fenton and TiO2 phototreatments, partially photodegraded solutions of 6 separate
pesticides (alachlor, atrazine, chlorfenvinphos, diuron, isoproturon and
pentachlorophenol) have been tested for biocompatibility, which have been evaluated
according to the Zahn-Wellens procedure. This study has been the topic of Chapter 5 part
A. In particular, it has been investigated if Microtox© could be considered as a suitable
global indicator of the evolution of biocompatibility of the contaminated water during the
phototreatment. The obtained results have demonstrated that biodegradability was
significantly increased after short photo-Fenton treatment times and that Microtox© acute
toxicity testing was useful to correctly represent overall dynamics and efficiency of
photo-treatment.
Therefore, it has been aimed to find out the best moment for coupling the photocatalytic
and biological processes. Thus, in Chapter 5 part B, the biotraitability of partially
photodegraded solutions of mixed pesticides has been more precisely investigated using
several packed-bed bioreactors, which have been successfully operated during 28 days.
178
The obtained concomitant decrease in toxicity and increase in biodegradability for the
partially phototreated solutions have confirmed that photo-Fenton was a promising pre-
treatment process, capable to enhance the biotreatability of waters contaminated with
biorecalcitrant chemicals (pesticides). Finally, more than 80% of DOC has been degraded
in the coupled system for pesticide solutions which had been photo-treated during only
0.6 h (t30W) with 1 g l-1 H2O2. This step of photo-Fenton treatment has been proved to be
the most relevant moment for coupling, because in this case, the biological treatment had
contributed to more than 50% of the total carbon conversion of the pesticide solution.
Recirculation of the bioreactor effluents has been presented as an outlook to improve the
global degradation degree of the contaminated waters.
In order to prospect for an industrial application of the integrated photocatalytic-
biological process, the development, installation and operation of an hybrid reactor has
been achieved in the framework of the CADOX project (European Commission). At
present, the testing of real industrial wastewaters is scheduled to provide interesting
information about the limiting factors due to realistic experimental conditions.
Finally, in the light of these studies, three main points have been clearly demonstrated:
o Biodegradability assessment is crucial to select the appropriate process for
removing problematic xenobiotics,
o Biological processes can be induced and enhanced for treating biodegradable
xenobiotics,
o Photocatalysis using TiO2 and photo-Fenton methods can be used as preliminary
treatment before a consecutive biological process for removing biorecalcitrant
xenobiotics.
179
Acknowledgments
This thesis has been carried out within the framework of the CADOX Project (Contract
N° EVK1-CT-2002-00122, European Commission) and EUREKA Project, “Bioreactor
For Innovative Mass Bacteria Culture, BIOMAC” (see: www.eureka.be, project E!2497).
181
Curriculum Vitae
Miléna LAPERTOT
Maupas 44, 1004 Lausanne
Phone: +41 21 646 53 16
Born on May, 15th, 1976 in Nîmes, France
EDUCATION
2003-2006 Docteur ès Sciences
EPFL, Ecole Polytechnique Fédérale de Lausanne – Switzerland
Thesis : A Strategy for Xenobiotic Removal using Photocatalytic
Treatment, Microbial Degradation or Integrated Photocatalytic-Biological
Process.
1999-2001 Postgraduated in Environmental Sciences
EPFL, Ecole Polytechnique Fédérale de Lausanne – Switzerland
1996-1999 Food and Nutrition Engineer
ENSBAN, Ecole Nationale Supérieure de Biologie Appliquée à la
Nutrition et à l’Alimentation – Dijon, France
PROFESSIONAL EXPERIENCE
2003-2006 Research assistant at the Laboratory of Environmental Biotechnology,
EPFL, – Switzerland
2001-2003 Scientific collaborator at the Laboratory of Environmental Biotechnology,
EPFL, – Switzerland
LANGUAGE French (mother tongue), English (fluent), Spanish (spoken knowledge)
COMPUTER SKILLS
Word, Excel, PowerPoint, LabVIEW, Photoshop, Endnote
OTHER ACTIVITIES
Fitness Trainer (8 hrs/week), personal development, manual activities
182
PUBLICATIONS
Lapertot, M., Ebrahimi, S., Dazio, S., Rubinelli, A., Pulgarin, C., 2006. Photo-Fenton and
biological integrated process for degradation of a mixture of pesticides. J. Photochem.
Photobiol. A., Submitted.
Lapertot, M., Pichat, P., Parra, S., Guillard, C., Pulgarin, C., 2006. Photocatalytic
degradation of p-halophenols in TiO2 aqueous suspensions: halogen effect on removal
rate, aromatic intermediates and toxicity variations. J. Environ. Sci. Health., A. In press,
Corrected Proof.
Lapertot, M., Pulgarín, C., 2006. Biodegradability Assessment of Several Priority
Hazardous Substances: Choice, Application and Relevance Regarding Toxicity and
Bacterial Activity. Chemosphere, In press, Corrected Proof.
Lapertot, M., Pulgarin, C., Fernández-Ibáñez, P., Maldonado, M. I., Pérez-Estrada, L.,
Oller, I., Gernjak, W., Malato, S., 2006. Enhancing biodegradability of priority
substances (pesticides) by solar photo-Fenton. Wat. Res. 40, 1086-1094.
Lapertot, M., Seignez, C., Ebrahimi, S., Peringer, P., 2006. Mass production of Bacterial
Communities Adapted to the Degradation of Volatile Organic Compounds (TEX)
Biodegradation, Submitted.
Lapertot, M., Seignez, C., Ebrahimi, S., Peringer, P. , 2006. Enhancing Production of
Adapted Bacteria to Degrade Chlorinated Aromatics. Ind.Eng.Chem.Res., In Press.
CONFERENCES
Lapertot, M., Domeniconi, P., Seignez, C., Ebrahimi, S., Peringer, P., 2005. Upgraded
conventional cultivation techniques for the mass production of bacterial communities
adapted to the degradation of volatile organic compounds (TEX). In Biotechniques for
Air Pollution Control, Proceedings of the International Congress Biotechniques for Air
Pollution Control. Coruña, Spain, October 5-7.
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